GUIDELINES FOR THE BIOREMEDIATION OF MARINE SHORELINES AND FRESHWATER WETLANDS

Prepared by:

1 Xueqing Z hu, 2 Albert D. Venosa, 1 Makram T. Suidan, and 3 Kenneth Lee

1 University of Cincinnati Department of Civil and Environmental Engineering Cincinnati, OH 45221

2 U.S. Environmental Protection Agency National Risk Management Research Laboratory Cincinnati, OH 45268

3 Department of Fisheries and Oceans-Canada Bedford Institute of Oceanography Marine Environmental Sciences Division Dartmouth, Nova Scotia B2Y 4A2

September, 2001

U.S. Environmental Protection Agency

Office of Research and Development National Risk Management Research Laboratory Land Remediation and Pollution Control Division 26 W. Martin Luther King Drive Cincinnati, OH 45268Acknow ledgments
The preparation of this guidance docum en t was perform e d under the direction of Albert D. Venosa of EPA’s Tr eatm e nt and Destruction Branch, Land Rem e diation and Pollution Control Division, National Risk Managem e nt Re search Laboratory, Cincinnati, OH. The report was prepared under EPA Contract68-C7-0057, Task Order 23 by the University of Cincinnati with the assistance of the Departm e nt of Fisheries and Oceans-Canada.

Disclaimer This report has been reviewed and approve d for publication by the Land Rem e diation and Pollution Control Division, National Risk Mana gem e nt Research Laboratory of the U.S. Environm ental Protection Agency. Mention of com p any nam e s, trade nam e s, or comm e rcial products does not constitute EPA endorsem e nt or recomm e ndations for use.

Further Information
For further inform ation, contact: Albert D. Venosa, Ph.D. U.S. EPA 26 W

. Martin Luther King Drive Cincinnati, OH 45268 Tel: 513-569-7668 Fax: 513-569-7105 Em ail: venosa.albert@epa.govEXECUTIVE SUMMARY The objective of this docum ent is to present a detailed technical guidance docum ent for use by spill responders for the biorem ediation of m arine shorelines and freshwater wetlands contam inated with oil and oil products. Techni cal personnel who are responsible for designing and operating field biorem ediation processes as we ll as consultants and equipm ent m a nufacturers will also f i nd it usef ul. This m a nual presents a rational approach f o r the design of biorem ediation processes pertinent to cleanup of oil-contam inat ed m a rine shorelines and freshwater wetlands. This docum ent evaluates current practices and st ate-of-the-art research results pertaining to biorem ediation of hydrocarbon contam ination re lative to types and am ounts of am endm ents used, frequency of application, assessm ent of the extent of biorem ediation, sam p ling, and analysis. The scope of the docum ent is lim ited to m a rine shorelines and freshwater wetlands because of definitive results from recently co m p leted, EPA-sponsored field studies. The final product is presented in a report form that is understandable by responders, on-scene coordinators, and rem e diation specialists. This report includes a thorough review and critique of the literature and theories pertinent to oil biodegradation, nutrient dynam i cs in shorelines, and analytical chem istry of oil and rem e diation nutrients. A planning approach to site identification, ev aluation, and selection along with inform ation on field investigations is also presented. The m anual includes exam ples of biorem ediation options and case studies of biorem ediation applied to m a rine shorelines and freshwater wetland environm ents.
The contents of this docum ent are arranged in a logical sequence first to provide basic inform ation for the evaluation of biorem ediation as a spill response option followed by guidelines for application that includes m e thods to m onitor its effectiveness. Thus, Chapter 1 presents an overall introduction and backgr ound discussion of biorem ediation including occurrence of oil spills, response m e thodologies, and a summ a ry of the scope, organization, and objective of the m a nual. Chapter 2 covers th e basic inform ation about oil, shorelines,
m echanism s of oil biodegradation, and a state- of-the-art review of controlled laboratory experim e nts and field trials of oil biodegr adation and nutrient dynam i cs in shoreline environm ents. For additional background info rm ation, Chapter 3 provides a m o re thorough review and critique of current analytical m e thods used to m onitor and verify oil spill biorem ediation success. Chapter 4 summ a rizes m a jor biostim ulatory and bioaugm enting
am endm ent m e thods and their application strategies

. Chapter 5 is the h eart of the docum ent and presents the actual guidelines for designing, planning, and im plem enting oil biorem ediation in the field, including site characterization, evalua tion of appropriate biorem ediation technologies, and the selection of the m o st appropriate t echnology for a specific site. Finally, Chapter 6 provides guidelines for assessm ent and interpretation of field results and provides help in assessing endpoints of biorem ediation (i.e., when treatm e nt is considered com p lete). The overall conclusions reached by the guidance m anual are as follows. First, with respect to
m a rine sandy shorelines, natural attenuation m a y be appropriate if background nutrient concentrations were high enough that intrinsic bi odegradation would take place at close to the expected m a xim u m rate. The Delaware study prove d this clearly. Certainly in nutrient-lim ited places like Prince W illiam Sound, Alaska, nutrient addition should accelerate cleanup ratesm a ny-fold. However, the decision to use the natural attenuation approach m a y be tem p ered by the need to protect a certain habitat or vital resource from the im pact of oil. For exam ple, using Delaware as the m odel, every spring season, horseshoe crabs m i grate to the shoreline of Delaware for their annual m a ting season. Millions of eggs are laid and buried a few mm below the surface of the sand. Migrating birds m a king thei r way from South Am erica to Arctic Canada fly by this area and feed upon these eggs to provide energy to continue their long flight. If an oil spill occurred in February or March, it would certainly be appropriate to institute biorem ediation to accelerate the disappearance of the oil prior to the horseshoe crab m a ting season despite the expected high natural attenuation rate. So, evenin the case where background nutrients are high enough to support rapid biodegradation, addition of m o re nutrients would help protect such a vital resource. If the spill occurred during the summ e r, and no vital natural resources were threatened by the spill, then reliance on natural attenuation m i ght be the wisest course of action. Of course, rem oval of free product and high con centrations of oil should still be conducted by conventional m eans even if a no biorem ediation action is warranted by the circum stances.
W ith respect to freshwater wetlands, the St . Lawrence River study dem onstrated that, if significant penetration of oil takes place into th e subsurface, biodegradation would take place very slowly and ineffectively. This is because of the anaerobic conditions that quickly occur in these types of saturated environm ents, and anaer obic biodegradation of petroleum oils is m u ch slower and less com p lete than under aerobic c onditions. One of the objectives of the St. Lawrence River experim e ntal design was to determ ine the am enability of wetlands to biodegradation when oil has penetrated into the sedim e nt. The oil was artif icially raked into the sedim e nt to m i m i c such an occurrence. Conse quently, no significant tr eatm e nt effects were observed because all the nutrients in the worl d would not stim ulate biodegradation if oxygen were the prim ary lim iting m a terial. If penetr ation did not take place beyond a few mm , then biorem ediation m i ght be an appropriate cleanup technology, since m o re oxygen would be available near the surface. It is clear that whatever oxygen gets transported to the root zone by the plants is only sufficient to support plant growth and insufficient to support the rhizosphere
m i croorganism s to degrade contam inating oil. However, if ecosystem restoration is the prim ar y goal rather than oil cleanup, the St. Lawrence River study strongly suggested that nutrient addition would accelerate and greatly enhance restoration of the site. Abundant plant growth t ook place in the nutrient-treated plots despite the lack of oil disappearance from the extra nutrients. Furtherm ore, the stim ulation lasted m o re than one growing season even though nutrients were ne ver added after the first year. Clearly, the plants took up and stored the extra nitrogen for us e in subsequent growing seasons, so restoration of the site was abundantly evident in a few short m onths. Thus, in conclusion, the decision to biorem ediate a site is dependent on cleanup, restoration, and habitat protection objectives and whatever factors that are present that would have an im pact on success. Responders m u st take into considerati on the oxygen and nutrient balance at the site. If the circum stances are such that no am ount of nutrients will accelerate biodegradation, then the decision should be m a de on the need to accelerat e oil disappearance to protect a vital living resource or sim p ly to speed up restoration of the ecosystem . If there is no imm e diate need to protect a vital resource or restore the ecosystem , then natural attenuation m a y be the appropriate response action. These decisions are clearly influenced by the circum stances of the spill.TABLE OF CONTENTS

Chapter 1 INTRODUCTION ..............................................................................................1 1.1 Occurrence of Oil Spills ..................................................................................1 1.2 Response to Oil Spills to Marine Shorelines and Freshwater Environm ents ..................................................................................3 1.2.1 Natural m e thods ..................................................................................3 1.2.2 Physical m e thods ..................................................................................4 1.2.3 Chem ical m e thods ..................................................................................5 1.3 Biorem ediation as an Oil Spills Cleanup Technology ..................................5 1.4 Scope of This Docum e nt ..................................................................................7 1.4.1 Objectives ..............................................................................................7 1.4.2 Organization of the guidance docum ent ..............................................7 Chapter 2 FACTORS AFFECTING NATURAL OIL BIODEGRADATION AND BIOREMEDIATION SUCCESS ..........................................................9 2.1 Physical-Chem i cal Properties of Crude Oil and Oil Products ......................9 2.1.1 Chem ical com position of crude oils and oil products ......................9 2.1.2 Physical properties of oil ....................................................................12 2.2 Behavior of Oil in the Environm ent ....................................................................13 2.2.1 W eathering processes ................................................................................13 2.2.2 Oil and shorelin e interactions ....................................................................16 2.2.3 Shoreline sensitivities ................................................................................16 2.3 Biodegradation of Oil ............................................................................................17 2.3.1 Mechanism of oil biodegradation: a m i crobiological perspective ........17 2.3.2 Environm ental factors affecting oil biodegradation ................................21 2.3.3 Evaluation of oil biodegradation: application of biom arkers ....................23 2.4 Laboratory Studies on Biorem ediation of Oil ............................................24 2.4.1 Bioaugm entation ................................................................................25 2.4.2 Biostim ulation ................................................................................26 2.5 Dem onstrations of Oil Biorem ediation Under Field Conditions ....................28 2.5.1 Mesocosm studies ................................................................................29 2.5.2 Field dem onstration ................................................................................30 2.5.3 Kinetics of oil biorem ediation ........................................................40 2.6 Nutrient Hydrodynam i cs ................................................................................41 2.6.1 Nutrient transport in beaches: a m e socosm study ................................42 2.6.2 Nutrient transport inbeaches: field trials ............................................43 Chapter 3 METHODS USED IN MONITORING OIL BIOREMEDIATION ........47 3.1 Analytical Methods ............................................................................................47 3.1.1 Microbiological analysis ....................................................................47 3.1.1.1 Enum eration of hydrocarbon-degrading m i croorganism s: culture based techniques ........................................................47 3.1.1.2 Culture-independent population/comm unity techniques ........48 3.1.2 Chem ical analysis of nutrients ........................................................503.1.3 Chem ical analysis of oil and oil constituents ............................................51 3.1.3.1 Total petroleum hydrocarbon (TPH) techniques ....................52 3.1.3.2 Analysis of specific oil constituents ............................................53 3.2 Biom arkers ........................................................................................................57 3.2.1 Comm only used biom arkers ....................................................................57 3.2.2 The effect of contam inant redistribution on observed rem e diation rates ........................................................61 3.3 Sam p ling in the Field ............................................................................................63 3.3.1 Sam p ling strategies ................................................................................64 3.3.2 Field sam p ling experiences ....................................................................65 3.4 Monitoring General Site Background Conditions ............................................66 3.5.1 Oxygen ............................................................................................67 3.5.2 pH ........................................................................................................67 3.5.3 Tem p erature ............................................................................................67 3.5.4 Salinity ............................................................................................68 3.5 Monitoring of Biological Im pacts ....................................................................68 3.5.1 Bioassessm ent ............................................................................................68 3.5.2 Bioassays ............................................................................................71 3.5.2.1 Benthic invertebrates ....................................................................72 3.5.2.2 Microtox ................................................................................72 3.5.2.3 Fish ............................................................................................73 3.5.3 Application of Bioassays to Assess Biorem ediation in Marine Environm ents ....................................................................74 Chapter 4 TYPES OF AMENDMENTS AND CONSIDERATIONS IN THEIR APPLICATIONS ....................................................................75
4.1 Nutrient Am endm ent ............................................................................................75 4.1.1 W a ter-soluble nutrients ....................................................................75 4.1.2 Granular Nutrients (slow-release) ........................................................76 4.1.3 Oleophilic Nutrients ................................................................................76 4.2 Microbial Am endm ents ................................................................................78 4.3 Plant Am endm ents (Phytorem ediation) ........................................................80 4.3.1 Mechanism s of phytorem ediation ........................................................80 4.3.2 Considerations in application of oil phytorem ediation ....................81 4.3.3 Applications in m a rine shoreline and freshwater wetlands ....................82 4.4 Oxygen Am endm ent ............................................................................................83 4.4.1 Tilling ........................................................................................................84 4.4.2 Forced aeration ................................................................................84 4.4.3 Chem ical m e thods ................................................................................84 Chapter 5 GUIDELINES FOR BIOREMED IATION OF MARINE SHORELINES AND FRESHWATER WETLANDS: DECISION-MAKING AND PLANNING ........................................................................................................86 5.1 Pre-treatm e nt Assessm ent ................................................................................88 5.1.1 Oil type and concentration ....................................................................885.1.2 Background nutrient content ....................................................................89 5.1.3 Type of shorelines ................................................................................90 5.1.4 Other factors ............................................................................................92 5.1.5 Summ a ry of pretreatm e nt assessm ent ........................................................92 5.2 Selection of Nutrient Products ....................................................................93 5.2.1 Nutrient selection based on efficacy and toxicity ................................93 5.2.2 Environm ental factors affecting nutrient selection ................................97 5.3 Determ ination of the Optim al Nutrient Loading and Application Strategy ........97 5.3.1 Concentration of nutrients needed for optim al biostim ulation ........97 5.3.2 Nutrient applica tion strategies ........................................................99 5.4 Sam p ling and Monitoring Plan ..................................................................101 5.4.1 Im portant variables ..............................................................................101 5.4.2 Statistical considerati ons in sam p ling plan ..........................................103 5.5 Considerations for Freshwater W e tland Biorem ediation ..............................104 5.5.1 Characteristics of freshwater wetlands ..........................................104 5.5.2 Biorem ediation strategies in freshwater wetlands ..............................106 Chapter 6 GUIDELINES FOR ASSESSMENT OF FIELD RESULTS AND TERMINATION OF TREATMENT ......................................................109 6.1 Assessm ent of Oil Biodegradation Efficacy ......................................................109 6.1.1 Verification of oil biodegradation ......................................................109 6.1.2 Assessm ent of physical rem oval ......................................................113 6.1.3 Operational endpoints based on oil biodegradation ..............................115 6.2 Environm ental Assessm ents ..............................................................................115 6.2.1 Operational guidance from environm ental assessm ents for treatm e nt application ..................................................................116 6.3 Case Study: Environm ental Assessm ent of Biorem ediation Treatm e nts in a Tidal Freshwater Marsh ..............................................................................117 6.3.1 Alterations in eco system structure ......................................................117 6.3.2 Alterations in ecosystem function ......................................................118 6.3.2.1 Microbial response ..................................................................118 6.3.2.2 Microtox solid phase test ......................................................119 6.3.2.3 Algal solid phase assay ......................................................119 6.3.2.4 Cladoceran survival test ......................................................121 6.3.2.5 Am phipod survival test ......................................................121 6.3.2.6 Gastropod survival/histopathology ..........................................122 6.3.2.7 Acute and chronic effects on fish ..........................................124 6.4 Ecotoxicological Tests for Risk Assessm ent ......................................................128 6.5 Ecotoxicological Tests to Identify Operational Endpoints ..............................129
REFERENCES ..................................................................................................................130Chapter 1 INTRODUCTION
1.1 Occurrence of Oil Spills Modern society continues to rely on the use of petroleum hydrocarbons for its energy needs. Despite recent technological advances, accidental spills of crude oil and its refined products occur on a frequent basis during routine operations of extraction, transportation, storage, refining and distribution. It is estim ated that between 1.7 and 8.8 m illion m e tric tons of oil are released into the world’s water every year (NAS, 1985), of which m o re than 90% is directly related to hum an activities including deliberate waste dispos al. Contrary to popul ar perception, only one eighth of the oil released into the aquatic envi ronm ent is from tanker accidents. It is also estim ated that about 30% of the spilled oil en ters freshwater system s (Cooney, 1984). These figures are rather uncertain and can vary grea tly from year to year depending on sources of estim ation and spill incidents. Table 1.1 summ a ri zes the num ber of oil spills and the quantities of release from 1970 to 1999 based on Interna tional Tanker Owners Pollution Federation’s oil spill database (ITOPF, 2001). These data include oil spills of over seven tons from tankers,
com b ined carriers, and barges. Although the data suggest a reduction in oil spills, the trend m a y only represent a tem porary downward fluctuation that is part of erratic cycling over the long term (Etkin and W e lch, 1997).
Marine shorelines are im portant public and ecologi cal resources that serve as a hom e to a variety of wildlife and provide public recreation. Mari ne oil spills, particularly large scale spill accidents, have posed great threats and cause extensive dam a ge to the m a rine coastal environm ents. For exam ple, the spill of 37,000 m e tric tons (11 m illion gallons) of North Slope crude oil into Prince W illiam Sound, Alaska, from the Exxon Valdez in 1989 led to the m o rtality of thousands of seabirds and m a rine m a mm a ls, a significant reduction in population of m a ny intertidal and subtidal organism s, and m any long term environm ental im pacts (Spies et al ., 1996). In 1996, the Sea Empress released approxim a tely 72,000 tons of Forties crude oil and 360 tons of heavy fuel oil at Milfor d Haven in South W a les and posed a considerable threat to local fisheries, wildlife and tourism (Edwards and Wh ite, 1999; Harris; 1997).
Com p ared to m a rine oil spills, inland oil spills have received m u ch less attention. However, freshwater spills are very comm on, with m o re than 2000 oil spills, on average, taking place each year in the inland waters of the continental United States (Owens et al ., 1993). Although freshwater spills tend to be of a sm aller volum e than their m a rine counterparts (Stalcup, et al ., 1997), they have a greater potential to endanger pub lic health and the environm ent because they often occur within populated areas and m a y dir ectly contam inate surface water and groundwater supplies. For exam ple, in 1988, an Ashland O il Com p any storage tank in Pittsburgh ruptured and spilled about 2,500 tons (750,000 gallons) of di esel oil in the Monongahela River, which contam inated drinking water intakes and led to downstream water shortages as far as 200 m iles (Miklaucic and Saseen, 1989). These catastrophic accidents, especially the Exxon Valdez spill, have increased public awareness about the risks involved in the storage and tr ansportation of oil and oil products and have prom pted m o re stringent regulations, such as the enactm e nt of the 1990 Oil Pollution Act by

1Congress. However, because oil is so widely us ed, despite all the precautions, it is alm o st certain that oil spills and leakage will continue to occur. Thus, it is essential that we have ef f e ctive counterm easures to deal with the problem

.
Table 1.1 Oil spills from tankers and other vessels into world water (adapted from ITOPF’s Oil Spill Database, http://www.itopf.com /stats.htm l )
Num b er of Oil Spills Year < 700 m e tric tons > 700 m e tric tons Quantities of Oil Spills (x 10 3 m e tric tons)

1970 1971 1972 1973 1974 1975 1976 1977 1978 1979 1980 1981 1982 1983 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998 1999 6 18 49 25 91 97 67 65 54 59 51 49 44 52 25 29 25 27 11 32 50 27 31 30 27 21 20 27 22 19 29 14 24 32 26 19 25 16 23 34 13

6 3 11

8 8 7 10 10 13 13

8 9 11

7 2 3 10

4 5 301 167 311 166 222 342 369 298 395 608 206

44 11 384

28 88 19 30 198 178

61 435 162 144 105 9 79 67 10 24

21.2 Response to Oil Spills in Marine Shorelines and Freshw ater Environments Strategies f o r cleaning up an oil spill are greatly af f ected by a variety of f

actors, such as the type of oil, the characteristics of th e spill site, and occasionally political considerations. A num ber of approaches and technologies have been developed for controlling oil spills in m a rine shorelines and freshwater environm ents. These m e thods have been reviewed and described extensively in a
num ber of technical docum ents, such as: Shoreline Countermeasure Manual (NOAA, 1992), Options for Minimizing Environmental Impacts of Freshwater Spill Response (NOAA and API, 1994), Understanding Oil Spills and Oil Spill Response (U.S. EPA, 1999), and Oil Spill Response in the Marine Environment (Doerffer, 1992). The m o st comm only used shoreline cleanup options (Table 1.2) are briefly described in the following text.
Table 1.2. Conventional Shoreline Clean-up Options Category of Response Options Exam ple Technology
Natural m e thod Natural attenuation Physical m e thod Boom ing S k i mmi n g Manual rem oval (W iping) Mechanical rem oval W a shing Sedim e nt relocation/Surf-washing Tilling In-situ burning Chem ical m e thod Dispersants Dem u lsif iers Solidif iers Surface film chem icals

1.2.1 Natural methods Natural attenuation or natural recovery is basi cally a no-action option that allows oil to be
rem oved and degraded by natural m eans. For som e spills, it is probably m o re cost-ef f e ctive and ecologically sound to leave an oil-contam inated s ite to recover naturally than to attem p t to intervene. Exam ples of such cases are spills at rem o te or inaccessible locations when natural
rem oval rates are fast, or spills at sensitive s ites where cleanup actions m a y cause m o re harm than good. It should also be noted that when na tural attenuation is used as a clean up m e thod, a
m onitoring program is still required to assess the perform a nce of natural attenuation. Major natural processes that result in the rem oval of oils include:

3• Evaporation : Evaporation is the m o st im portant na tural cleansing process during the early stages of an oil spill, and it results in the rem oval of lighter-weight com ponents in oil. Depending on the com position of the oil spilled, up to 50 percent of the m o re toxic, lighter weight com ponents of an oil m a y evaporate within the first 12 hours following a spill (U.S. EPA, 1999).

• Photooxidation : Photooxidation occurs when oxyge n under sunlight reacts with oil
com ponents. Photooxidation leads to the breakdown of m o re com p lex com pounds into sim p ler com pounds that tend to be lighter in weight and m o re soluble in water, allowing them to be rem oved further through other processes.

• Biodegradation : Various types of m i croorganism s th at are capable of oxidizing petroleum hydrocarbons are widespread in nature. Bi odegradation is a particularly im portant m echanism for rem oving the non-volatile com ponents of oil from the environm ent. This is a relatively slow process and m a y require m onths to years for m i croorganism s to degrade a significant fraction of an oil stranded within the sedim e nts of m a rine and/or freshwater environm ents.
1.2.2 Physical methods Physical containm ent and recovery of bulk or fr ee oil is the prim ary response option of choice in the United States for the cleanup of oil spills in m a rine and freshwater shoreline environm ents. Comm only used physical m e thods include:

• Boom ing and skimm i ng : Use of boom s to contain and control the m ovem e nt of floating oil and use of skimm e rs to recover it. The envir onm ental im pact of this m e thod is m i nim a l if traffic of the cleanup work force is controlled.

• W i ping with absorbent m a terials : Use of hydrophobic m a terials to wipe up oil from the contam inated surface. Wh ile the disposal of contam inated waste is an issue, the environm ental effect of this m e thod is also lim ited if traffic of cleanup crew and waste generation is controlled.

• Mechanical rem oval : Collection and rem oval of oiled surface sedim e nts by using
m echanical equipm ent. This m e thod should be used only when lim ited am ounts of oiled
m a terials have to be rem oved. It should not be considered for cleanup of sensitive habitats or where beach erosion m a y result.

• Wa s h i n g : washing of the oil adhering along the shorelines to the water’s edge for collection. W a shing strategies range from low-pressure co ld water flushing to high-pressure hot water flushing. This m e thod, especially using high-pr essure or hot water, should be avoided for wetlands or other sensitive habitats.

• Sedim e nt relocation and tilling : Movem e nt of oiled sedim e nt from one section of the beach to another or tilling and m i xing the contam inat ed sedim e nt to enhance natural cleansing

4processes by f acilitating the dispersion of oil into the water columnand prom oting the interaction between oil and m i neral f i nes. Tilling m a y cause oil penetration deep into the shoreline sedim e nts. The potential environm enta l im pacts f r om the release of oil and oiled sedim e nt into adjacent water bodies should also be considered.

• In-situ burning : Oil on the shoreline is burned usually when it is on a com bustible substrate such as vegetation, logs, and other debris. This m e thod m a y cause significant air pollution and destruction of plants and anim als.
1.2.3 Chemical methods
Chem ical m e thods, particularly dispersants, have been routinely used in m a ny countries as a response option. For som e countries, suchas the United Kingdom , where rough coastal conditions m a y m a ke m echanical response probl em atic, dispersants are the prim ary choice (Lessard and Dem a rco, 2000). However, chem ical m e thods have not been extensively used in the United States due to the disagreem ent about their effectiveness and the concerns of their toxicity and long-term environm ental effects (U .S. EPA, 1999). Major existing chem ical agents include:

• Dispersants : dispersing agents, which contain surfactants, are used to rem ove floating oil from the water surface to disperse it into th e water columnbefore the oil reaches and contam inates the shoreline. This is done to reduce toxicity effects by dilution to benign concentrations and accelerate oil biodegradation rates by increasing its effective surface area.

• Dem u lsif iers : Used to break oil-in-water em ulsions and to enhance natural dispersion.

• Solidif iers : Chem icals that enhance the polym erization of oil can be used to stabilize the oil, to m i nim i ze spreading, and to increase the effectiveness of physical recovery operations.

• Surface film chem icals : Film -form i ng agents can be used to prevent oil from adhering to shoreline substrates and to enhance the rem ova l of oil adhering to surfaces in pressure washing operations.
1.3 Bioremediation as an Oil Spill Cleanup Technology Although conventional m e thods, such as physical rem oval, are the first response option, they rarely achieve com p lete cleanup of oil spills

. According to the Office of Technology Assessm ent (OTA, 1990), current m echanical m e thods typically recover no m o re than 10-15 percent of the oil af ter a m a jor spill. Biorem ediation has em erged as one of the m o st prom ising secondary treatm e nt options for oil rem oval since its successful application after the 1989 Exxon Valdez spill (Bragg et al ., 1994; Prince et. al ., 1994). Biorem ediation has been defined as “the act of adding m a terials to contam inated environm ents to cause an acceleration of the natural biodegradation processes” (OTA, 1991). This t echnology is based on the prem ise that a large percentage of oil com ponents are readily biode gradable in nature (Atlas, 1984, 1981; Prince, 1993). The success of oil spill biorem ediation depe nds on our ability to establish and m a intain

5conditions that favor enhanced oil biodegradation rates in the contam inated environm ent. There are two m a in approaches to oil spill biorem ediation:

• bioaugmentation, in which known oil-degrading bacteria are added to supplem ent the existing m i crobial population, and

• biostimulation , in which the growth of indigenous oil degraders is stim ulated by the addition of nutrients or other growth-lim iting cosubstr ates, and/or by alterations in environm ental conditions (e.g. surf-washing, oxygen addition by plant growth, etc.). Both laboratory studies and field tests have shown that biorem ediation, biostim ulation in particular, can enhance oil biodegradation on cont am inated shorelines (Prince, 1993; Swannell, et al. , 1996). Recent field studies have also dem onstrat ed that biostim ulation is a m o re effective approach because the addition of hydrocarbon de grading m i croorganism s will not enhance oil degradation m o re than sim p le nutrient addition (Lee et al , 1997a; Venosa et al ., 1996; see Chapter 2 in detail). Biorem ediation has seve ral advantages over conventional technologies. First, the application of biorem ediation is re latively inexpensive. For exam ple, during the cleanup of the Exxon Valdez spill, the cost of biorem ediating 120 km of shoreline was less than one day’s costs for physical washing (Atla s, 1995). Biorem ediation is also a m o re environm entally benign technology since it involves the eventual degradation of oil to m i neral products (such as carbon dioxide and water), wh ile physical and chem ical m e thods typically transfer the contam inant from one environm ental com p artm ent to another. Since it is based on natural processes and is less intrusive and disruptive to the contam inated site, this “green technology” m a y also be m o re acceptable to the general public.
Biorem ediation like other technologies also has its lim itations. Biorem ediation involves highly heterogeneous and com p lex processes. The su ccess of oil biorem ediation depends on having the appropriate m i croorganism s in place under suitabl e environm ental conditions. Its operational use can be lim ited by the com position of the oil sp illed. Biorem ediation is also a relatively slow process, requiring weeks to m onths to take eff ect, which m a y not be feasible when imm e diate cleanup is dem a nded. Concerns also arise about potential adverse effects associated with the application of biorem ediation agents. These in clude the toxicity of biorem ediation agents them selves and m e tabolic by-products of o il degradation and possible eutrophic effects associated with nutrient enrichm e nt (Swannell et al ., 1996). Biorem ediation has been proven to be a cost-effective treatm e nt tool, if used properly, in cleaning certain oil-contam inated environm ents. Few detrim ental treatm e nt effect s have been observed in actual field operations. Currently, one of the m a jor challenges in the app lication of oil biorem ediation is the lack of guidelines regarding when and how to use this technology. Although extensive research has been conducted on oil biorem ediation during the last decade, m o st existing studies have concentrated on either evaluating the f easib ility of biorem ediation f o r dealing with oil contam ination, or testing favored products and m e thods (Mearns, 1997). Only lim ited num ber of pilot-scale and field trials, which m a y provide the m o st convincing dem onstrations of this technology, have been carried out. To m a ke m a tters worse, m a ny field tests have not been properly designed, well controlled or correctly an alyzed, leading to skepticism and confusion

6am ong the user comm unity (Venosa, 1998). There is an urgent need for a detailed and workable set of guidelines for the application of this technology for oil spill responders that answers questions such as when to use biorem ediation, what biorem ediation agents should be used, how to apply them , and how to m onitor and evaluate th e results. Scientific data for the support of an operational guidelines docum ent has recently been provided from laboratory studies and fields trials carried out by the U.S. Environm ental Prot ection Agency (EPA), University of Cincinnati, and Fisheries and Oceans Canada.
1.4 Scope of This Document
1.4.1 Objectives The objective of this m a nual is to provide detaile d and applicable technical guidelines for use by spill responders for the biorem ediation of m a rine shorelines and freshwater wetlands that are contam inated with crude oil and its refined pr oducts. The docum ent will also provide guidance to scientists, other technical personnel, and m a nufac turers with an interest in the design and im plem entation of field biorem ediation processes. The docum ent evaluates current practices and state-of-the-art research results pertaining to biorem ediation of hydrocarbon contam ination and presents a rational procedure for the design of bi orem ediation processes pertinent to clean-up of oil stranded within sedim e nts of shoreline environm ents. The scope of this docum ent is lim ited to m a rine shorelines and freshwater wetlands where the effectiveness of biorem ediation has been quantif ied in controlled field and case studies following spill incidents. To date, there is no conclusive evidence of successful oil spill clean up operations in the open sea by the application of biorem ediation strategies (Atlas, 1995; Swannell et al ., 1996). W ith respect to freshwater environm en ts, an oil spill is m o st likely to have the greatest im pact on sensitive wetlands or m a rshes ra ther than running rivers. To provide insight into the f easibility of biorem ediation strategies in these habitats, the inf o rm ation presented in this
m a nual is based on experim e ntal oil spills conducted in the 1990s jointly by EPA, the University of Cincinnati, and Fisheries and Oceans-Canada . The first study was a field experim e nt on a sandy beach in Delaware in 1994. In 1999 and 2000, a sim ilar field investigation was conducted on a freshwater wetland along the shoreline of the St. Lawrence River, Quebec, Canada. Another field study was conduced in Dartm outh, Nova Scotia in 2000 on a salt m a rsh dom inated by Spartina alterniflora . This docum ent will provide oil biorem ediation guidelines based on current practices and research with the em phasis on the findings of these fields studies. Because the study conducted on the shoreline of Nova Scotia has not been concluded at the tim e of this writing, guidelines of oil biorem ediation in salt m a rshes will be available as a supplem entary docum ent upon the conclusion of this investigation.
1.4.2 Organiz ation of the guidance document For ease of use, the contents of this docum ent are arranged in a logical sequence first to provide basic inform ation for the evaluation of biorem ediation as a spill response option followed by guidelines for application that includes m e thods to m onitor its effectiveness. Thus, Chapter 2 covers the basic inform ation about oil, shorelines, m echanism s of oil biodegradation, and a state7of-art review of controlled laboratory experim e nts and field trials of oil biodegradation and nutrient dynam i cs in shoreline environm ents. For additional background inform ation, Chapter 3 provides a m o re thorough review and critique of current analytical m e thods used to m onitor and verify oil spill biorem ediation success. Ch apter 4 summ a rizes m a jor biostim ulatory and bioaugm enting am endm ent m e thods and their app lication strategies. Chapter 5 presents guidelines for designing, planning, and im plem enti ng oil biorem ediation in the field, including site characterization, evaluation of appropriate biorem ediation technologies, and the selection of the m o st appropriate technology for a specific s ite. Finally, Chapter 6 provides guidelines for
assessm ent and interpretation of f

i eld results.

8Chapter 2 FACTORS AFFECTING NATURAL OIL BIODEGRADATION AND BIOREMEDIATION SUCCESS

Oil biorem ediation is a com p lex process involvi ng interactions of oil and m i croorganism s under the conditions of the prevailing environm ent. To understand the scope and strategies of oil biorem ediation, it is essential to first understand th e properties of oil, the environm ent of concern (e.g., m a rine shorelines and freshwater wetlands), the fate of oil in that environm ent, the
m echanism s of oil biodegradation and the factors that control its rate.
2.1 Physical-Chemical Properties of Crude Oil and Oil Products
Crude oil and petroleum products are very com p lex and variable m i xtures of thousands of individual com pounds that exhibit a wide range of physical properties. Understanding these properties is im portant in determ ining behavior of spilled oil and the appropriate response option. The com position and properties of various petrol eum hydrocarbons have been described in detail by Clark and Brown (1977) and the National Academy of Sciences (1985). Large oil property databases also exist such as the one pos ted on the Internet by Environm ent Canada ( www.etcenttre.org/spills ), which contains inform ation on over 400 oils (Jokuty, et al ., 2000).
2.1.1 Chemical composition of crude oils and oil products
Crude Oil Crude oil is com p rised of both hydrocarbon com pounds (accounting for 50–98% of total
com position) and non-hydrocarbon com pounds (cont aining sulfur, nitrogen, oxygen, and various trace m e tals) in a wide array of com b inati ons (Clark and Brown, 1977). The chem ical
com position and physical characteristics of severa l crude oils is illustrated in Table 2.1. Som e representative organic com pounds found in crude oil are illustrated in Figure 2.1. Petroleum
com ponents m a y be classified into four m a jor groups based on their differential solubility in organic solvents (Leahy and Colwell, 1990).
1. Saturated hydrocarbons: Include norm a l and branched alkanes with structures of C n H 2n+2 (aliphatics) and cyclic alkanes with structures of C n H 2n (alicyclics), which range in chain length from one carbon to over 40 carbons. Sa turates usually are the m o st abundant constituents in crude oils.
2. Arom atic hydrocarbons: Include m onocyclic ar om atics (e.g., benzene, toluene, and xylenes) and polycyclic arom atic hydrocarbons (PAHs) (e.g., naphthalene, anthracene, and phenanthrene), which have two or m o re fused arom atic rings. PAHs are of particular environm ental concern because they are poten tial carcinogens or m a y be transform e d into carcinogens by m i crobial m e tabolism

.

93. Resins: Include polar com pounds containi ng nitrogen, sulfur, and oxygen (e.g., pyridines and thiophenes). They are often referred to as NSO com pounds.
4. Asphaltenes: Consist of poorly characterized high m o lecular weight com pounds that include both high m o lecular weight and poorly character ized hydrocarbons and NSOs. Metals such as nickel, vanadium , and iron are also associated with asphaltenes.

Table 2.1 Chem ical com position and physical pr operties of representative crude oils (adapted from Clark and Brown, 1977) Characteristic or Com ponent Prudhoe Bay South Louisiana Kuwait

API gravity (20 o C) Sulfur (wt%) Nitrogen (wt%) Nickel (ppm ) Vanadium (ppm ) Naphtha fraction (wt%) a Saturates Arom atics Resins & Asphaltenes High-boiling fraction (wt%) b Saturates Arom atics Resins & Asphaltenes 27.8 34.5 31.4 0.94 0.25 2.44 0.23 0.69 0.14 10 2.2 7.7 20 1.9 28
23.2 18.6 28.0 19.9 16.5 20.3 3.2 2.1 2.4

-- -- -- 76.8 81.4 77.3 47.7 56.3 34.0 25.0 16.5 21.9 4.1 8.6 21.4 These analyses represent values for one typical cr ude oil from three distinct geographical regions; variations in com position can be expected for oils produced from different form ations or fields within each region.
a Fraction boiling from 20 o to 205 o C b Fraction boiling above 205 o C

10S A TUR A T ES AR O M ATI C S

H 3 CC C C C CH 3 H 2 H 2 H 2 H 2

n- he x a ne

n- he p ta d e c a n e ( n- C 17

H 36 )

pr i s t a n e ( C 19 H

40 )

H H

17 α (H ), 2 1 β ( H )

- h opa n e

RE S I NS

N CH 3 2- m e t h yl p y r i d in e

di

b e nzoth io p h e n e S CH 3

to lu e n e

n

a

ph t h al

en e

ch r y s e n e

be nz o [ a ] p y r e n e

A S P H AL TE NE S

(C 79 H 92 N

2 S 2 O) 3

F i gur e 2.1 R e pr es e n ta t iv e o r g a ni c c o m p o un d s f o un d in c r u d e o i l s

11Refined oil products Refined petroleum products, such as gasoline, kerose ne, jet fuels, fuel oils, and lubricating oils, are derived from crude oils through processe s such as catalytic cracking and fractional distillation. These products have physical and ch em ical characteristics that differ according to the type of crude oil and subsequent refining pr ocesses. They contain com ponents of crude oil covering a narrow range of boiling points. Inaddition, during catalytic cracking operations, unsaturated com pounds, or olefins (alkenes and cy cloalkenes), which are not present in crude oils, can be form ed. The concentrations of olef ins are as high as 30% in gasoline and about 1% in jet fuel (NAS, 1985). A list of chem ical com positions of the fractions of crude oils and the refined products is shown in Table 2.2.
Table 2.2 Chem ical com positions of refined petroleum products (adapted from Clark and Brown, 1977) Distillation Hydrocarbon Range of Typical Refined Fraction Types Carbon Atom s Products Gasoline & naphtha Saturates 4-12 Gasoline Olefins Arom atics
Middle distillate Saturates 10-20 Kerosene Olefins Jet fuel Arom atics Heating oils Diesel oils W i de-cut gas oil Saturates 18-45 Wa x Arom atics Lubricating oil
Residum Resins >40 Residual oils Asphaltenes Asphalt

2.1.2 Physical properties of oil
Im portant physical properties of oil that affect its behavior in the environm ent and spill cleanup responses include:
1. Density: Two types of density expressions fo r oils are often used: specific gravity and Am erican Petroleum Institute (API) gravity. Sp ecif i c gravity is the ratio of the m a ss of a substance to the m a ss of the equivalent volum e of water at a specified tem p erature. The API gravity arbitrarily assigns a value of 10° to pure water at 10°C (60°F). The API gravity can be calculated from the specific gravity using the form ula:

12o 141

. 5 API Gravity 

# ( ) = Specific Gavity ( 16 o C / 60 o F ) − 131

. 5 ( 2

. 1 )

Oils with low densities or low specif ygravitie s have high API gravities. Crude oils have specific gravities in the range of 0.79 to 1.00 (equi valent to API Gravities of 10 to 48) (Clark and Brown, 1977). Oil density is an im portant i ndex of oil com position that is frequently used to predict its f a te in water.
2. Viscosity: Viscosity is the property of a fluid that describes how it resists a change in shape
or m ovem e nt. The lower the visc osity a fluid has, the m o re easily it flows. The viscosity of petroleum is related to oil com positions and th e am bient tem p erature. It is an im portant index of the spreading rate of a spilled oil.
3. Pour Point: The pour point of an oil is the tem p erature at which it becom e s sem i -solid or stops flowing. The pour point of crude oils varies from –57°C to 32°C. It is another im portant characteristic with respect to oil fate and cleanup strategies.
4. Solubility in water: The solubility of oil in water is extrem ely low and depends on the chem ical com position of the petroleum hydro carbon in question and tem p erature. For a typical crude oil, solubility is around 30 m g /L (NAS, 1985). The m o st soluble oil
com ponents are the low m o lecular weight arom a tics such as benzene, toluene and xylene. This property is im portant with respect to oil fate, oil toxicity and biorem ediation processes.
Other im portant physical properties of oils incl ude flash point, vapor pressure, surface tension, and adhesion.
2.2 Behavior of Oil in the Environment 2.2.1 Weathering processes
Wh en oil is introduced into the environm ent, it imm e diately goes through a variety of physical, chem ical and biological changes (Figure 2.2). These weathering processes will alter oil
com position and properties in ways that m a y aff ect spill response strategies. Biorem ediation is a relatively slow process that is often used as a polishing step after conventional cleanup options have been applied. Thus the residual target oil m a y be extensively weathered prior to the deploym ent of biorem ediation strategies.
W eathering processes, including biodegradation, ha ve been reviewed and described extensively in the literature (Clark and MacLeod, 1977; Jordan, R.E. and Payne, J.R., 1980; National Academy of Sciences, 1985). Major physical and chem ical fates of oil are briefly summ a rized in this section and the biological fate will be discussed in section 2.3.

13Spreading The spreading of oil on water is one of the m o st im portant processes during the first hours of a spill, provided that the oil pour point is lower than the am bient tem p erature. The principal forces influencing the spreading of oil include gravity, inertia, friction, viscosity and surface tension. This process increases the overall surface area of the spill, thus enhancing m a ss transfer via evaporation, dissolution, and later biodegradation. Evaporation In term s of environm ental im pacts, evaporation is the m o st im portant weathering process during the early stages of an oil spill in that it can be responsible for the rem oval of a large fraction of the oil including the m o re toxic, lower m o lecu lar weight com ponents. For oil on water, evaporation rem oves virtually all the norm a l alkanes sm aller than C 15 within 1 to 10 days. Volatile arom atic com pounds, such as benzene and to luene, can also be rapidly rem oved from an oil slick through evaporation. However, these vol atile oil com ponents m a y be m o re persistent when oil is stranded in sedim e nts. The volatile com ponents m a ke up 20-50% of m o st crude oils, about 75% of No. 2 fuel oil, and about 100% of gasoline and kerosene. As a result, the physical properties of the rem a ining slick change significantly (e.g., increased density and viscosity). Major factors influencing the rate of evapora tion include com position and physical properties of the oil, wave action, wind velocity, and water tem p erature (Clark and MacLeod, 1977; Jordan, R.E. and Payne, J.R., 1980).
Dissolution Although dissolution is less im portant from the viewpoint of m a ss loss during an oil spill, dissolved hydrocarbon concentrations in water are particularly im portant due to their potential influence on the success of biorem ediation and the e ffect of toxicity on bi ological system s. The extent of dissolution depends on the solubility of the spilled oil, weather conditions, and the characteristics of the spill site. The low m o lecu lar weight arom atics are the m o st soluble oil
com ponents, and they are also the m o st toxic com ponents in crude and refined oils. Although
m a ny of them m a y be rem oved through evaporati on, their im pact on the environm ent is m u ch greater than sim p le m a ss balance considerati ons would im ply (NAS, 1985). Dissolution rates are also influenced by photochem i cal and biological processes.
Photooxidation Photooxidation is another weathering process that m a y have im portant biological consequences. In the presence of oxygen, natural sunlight ha s sufficient energy to transform m a ny com p lex petroleum com pounds such as high m o lecular weight arom atics and polar com pounds into sim p ler com pounds through a series of free-radical chain reactions. This process m a y increase the solubility of oil in water, due to the form ation of polar com pounds such as hydroperoxides, aldehydes, ketones, phenols, and carboxylic acids. Detrim ental effects m a y be associated with this increase in the solubility of oil in water (i.e., bioavailability) and the f o rm ation of toxic
com pounds m e diated by photooxidation. On the other hand, the form ation of polar com pounds
m a y increase the rate of biodegradation of petroleum , particularly at lower concentrations where acute toxicity ef f ects are lim ited (Nicodem et al . 1997).

14Dispersion Dispersion, or form ation of oil-in-water em ulsi ons, involves incorporating sm all droplets of oil into the water column , resulting in an increase in surface area of the oil. Ingeneral, oil-in-water
em ulsions are not stable. However, they can be m a intained by continuous agitation, interaction with suspended particulates, and the addition of chem ical dispersants. Dispersion m a y influence oil biodegradation rates by increasing the cont act between oil and m i croorganism s and/or by increasing the dissolution rates of the m o re soluble oil com ponents. Emulsification The process of em ulsification of o ils involves a change of state from an oil-on-water slick or an oil-in-water dispersion to a water-in-oil em ulsi on, with the eventual possible form ation of a thick, sticky m i xture that m a y contain up to 80% water, comm only called “chocolate m ousse”. The f o rm ation and stability of em ulsions are prim arily related to the chem ical com position of the oils and are enhanced by wax and asphaltic m a terials. Surface-active m a terials generated through photochem i cal and biological processes are al so involved in form ation of the em ulsions. The form ation of em ulsions m a kes oil clean- up operations m o re difficult by decreasing the effectiveness of physical oil spill recovery pro cedures and suppressing the natural rates of oil biodegradation.
Other im portant physical and chem ical weatheri ng processes that influence the rates of oil degradation include adsorption onto suspended particulate m a terials, sinking and sedim e ntation, and tar ball form ation.

Phot oox i d at i o n E v apor at i o n

E m ul s i f i cat i o n

S p r eadi n g

D i sp e r sio n

Si nk i n g & Sedi m e n t at i o n A d s o r p t i on and P e n e tr a t io n
Mi g r a t i o n and R e l eas e Di s s o l u t i o n

B i odegr adat i o n B i odegr adat i o n
F i n e P a rt ic l e In t e ra c t io n

Fi gur e 2. 2 M a j o r w eat her i ng pr ocess e s af t e r an oi l spi l l

152.2.2 Oil and shoreline interactions
Wh en oil spills occur in m a rine or f r eshwater shoreline (e.g., f r eshwater wetlands) environm ents, interactions between the spilled oil and the shore f urther com p licate the weathering processes. It is very im portant to understand these interactions in determ ining the scope and lim itations of oil spill biorem ediation.
The behavior of spilled oil in shoreline environm ents is prim arily dependent on the properties of the shoreline, such as the porosity of the s ubstrate and the energy of the waves acting on a shoreline. Higher wave exposure enhances both physical rem oval and weathering processes. W a ve-swept rocky shores tend to recover f r om oil spills within a m a tter of m onths whereas
m a ngroves and m a rshes m a y act as a petroleum si nk for m a ny years. However, tidal pum ping is also a factor prom oting oil penetration into the se dim e nts. The rate and depth of oil penetration depend prim arily on the porosity of the substrate. On coarse-grained shorelines like cobble and sandy beaches, oil can penetrate deeper and rem a inlonger (when it is trapped below the lim it of wave action), com p ared to f i ner grained sedim e nts such as silts and clay. However, oil is m o re easily rem oved by water flushing from coarse-grained sedim e nts. Interactions of oil with tidal action, waves, and shoreline substrate m a y also f o rm asphalt-like oil-sedim e nt m a ts that are resistant to further biological and photochem i cal weathering.
Recent studies have shown that the interactions be tween oil and fine m i neral particles also play an im portant role in natural oil cleansing in m a rine shorelines (Bragg and Owens, 1995; Lee et al., 1997b). This process of oil and fine-particle interaction reduces the adhesion of oil to intertidal shoreline substrates through the form a tion of oil-m i neral fine flocs that are easily dispersed by tidal action and currents. More im portantly, oil-m i neral fine flocs enhance the availability of oil for biodegradation, and thus oil biodegradation rates are accelerated by this process (Lee et al ., 1997c).
2.2.3 Shoreline sensitivities
Marine shorelines and f r eshwater environm ents have a wide range of sensitivities to oil and clean-up activities. The National Oceanic & Atm o spheric Adm i nistration and the Am erican Petroleum Institute have developed the Environm ental Sensitivity Index (ESI) to classif y shoreline types for spill response (NOAA, 1992; NOAA and API, 1994). This classification schem e (e.g., Table 2.3) has been used in oil spill contingency planning and spill response operations (Hayes et al ., 1995). Major factors considered in ranking habitat sensitivity include shoreline type (substrate, grain size, tidal el evation), exposure to wave and tidal energy, biological productivity and sensitivity, and ease of cleanup. Biorem ediation m a y be effective and cause the least dam a ge on both the m oderately and the m o st sensitive shoreline types.
The Environm ental Sensitivity Index f o r f r eshwater shorelines is shown in Table 2.4, based on NOAA & API (1994) and Hayes et al . (1995 & 1997). Major factors considered in ranking ESI for these habitats include degree of exposure to natural rem oval processes, biological

16productivity and sensitivity, hum an use of the habita t, and ease of oil rem oval. Biorem ediation
m a y be f easible and cause the least dam a ge on both the m oderately sensitive and the m o st sensitive shoreline types, although its effectiveness is still uncertain due to the lack of sufficient research and field efficacy dem onstrations.
Table 2.3 Shoreline ESI ranking f o r habitats in m a rine shorelines (where 1 is least sensitive and 10 is m o st sensitive to oil and clean up actions) Environm ental Sensitivity Index (ESI) Shoreline Type 1
2 3 4 5 6 7 8 9 10 Exposed rocky shores Sea walls and piers Exposed wave-cut platform s Fine-grained sand beaches Coarse-grained sand beaches Mixed sand and gravel beaches Gravel beaches and riprap Exposed tidal flats Sheltered rocky shores Sheltered tidal f l ats Salt m a rshes and Mangroves

2.3 Biodegradation of Oil Biodegradation of oil is one of the m o st im portant processes involved in weathering and the eventual rem oval of petroleum from the environm ent, particularly for the nonvolatile
com ponents of petroleum

. Num e rous scientific review articles have covered various aspects of this process an d the environm ental factors thatinfluence the rate of biodegradation (Zobell, 1946 & 1973; Atlas, 1981 & 1984; NAS, 1985; Focht and W e stlake, 1987; Leahy and Colwell, 1990).
2.3.1 Mechanism of oil biodegradation: a microbiological perspective Distribution of hydrocarbon-degrading microorganisms Microorganism s capable of degrading petr oleum hydrocarbons and related com pounds are ubiquitous in m a rine, freshwater, and soil habita ts. Over 200 species of bacteria, yeasts and fungi have been shown to degrade hydrocarbons ranging from m e thane to com pounds of over 40 carbon atom s (Zobell, 1973). In the m a rine environm ent, bacteria are considered to be the predom inant hydrocarbon-degraders with a distribu tion range that even covers extrem e cold Antarctic and Arctic environm ents (Floodgate, 1984; Jordan and Payne, 1980). In the freshwater environm ent, yeast and fungi m a y also playa significant role in degrading petroleum hydrocarbons (Cooney, 1984). Som e of the m o st im portant hydrocarbon-degrading
m i croorganism s in both m a rine and freshwater environm ents are listed in Table 2.5.

17Table 2.4 Shoreline ESI ranking for habitats in fr eshwater shorelines (where 1 is least sensitive and 10 is m o st sensitive to oil and clean up actions) ESI Lacustrine aLarge Rivers a

1 Exposed rocky cliffs Exposed rocky banks Exposed m an-m a de structures Vertical, solid revetm ents banks Quiet pools with low-sensitive

2 Shelving bedrock shores Rocky shoals, bedrock Sm all, nonnavigable channel Ledges with m oderate currents and low-sensitive banks
3 Eroding scarps in Exposed, eroding banks in Navigable channel with unconsolidated sedim ents Unconsolidated sedim ents m oderate currents and lowsensitive banks
4 Sand beaches Sandy bars and gently Sm all, nonnavigable channel Sloping banks with rapids over bedrock
5 Mixed sand and gravel Mixed sand and gravel bars Navigable channel with rapids beaches over bedrock
6 Gravel beaches and riprap Gravel bars and riprap Channel with associated lowvulnerable upper bottom l and hardwoods
7 Exposed flats Not present Navigable stream s with associated wide swam ps on one side
8 Sheltered rocky shores Vegetated, steeply sloping Navigable stream s with Sheltered m an-m a de Bluffs associated wide swam ps on structures Sheltered m an-m a de both sides Structures
9 Sheltered vegetated low Vegetated low banks Meandering channel with banks Muddy substrates abundant leakage points into (unvegetated) associated swamps and oxbows
10 Sheltered sand flats Freshwater m arshes and Navigable anastom osing Sm all Rivers b

channel with abundant leakage points into associated swam ps Freshwater m arshes and

swam ps Swam ps

a ESI adapted from Hayes et al. (1995) b RSI (Reach Sensitivity Index) adapted from Hayes et al. (1997)

18Table 2.5 Representative m i croorganism s capable of degrading petroleum hydrocarbons (Based on Atlas, 1984; Focht and W e stlake, 1987; Jordan and Payne, 1980; Leahy and Colwell, 1990) Bacteria Yeast and Fungi
Achromobacter Acinetobacter Alcaligenes Arthrobacter Bacillus Brevibacterium Cornybacterium Flavobacterium Nocardia Pseudimonas Vibrio Aspergillus Candida Cladosporium Penicillium Rhodotorula Sporobolomyces Trichoderma

The distribution of hydrocarbon-utilizing m i croorga nism s is also related to the historical exposure of the environm ent to hydrocarbons. Thos e environm ents with a recent or chronic oil contam ination will have a higher percentage of hydrocarbon degraders than unpolluted areas. In “pristine” ecosystem s, hydrocarbon utilizers m a y m a ke up less than 0.1% of the m i crobial
comm unity; and in oil-polluted environm ents, th ey can constitute up to 100% of the viable
m i croorganism s (Atlas, 1981). It should be noted that there is no single strain of bacteria with the m e tabolic capacity to degrade all the com ponents found within crude oil. In nature, biodegradation of a crude oil typically involves a succession of species within the cons ortia of m i crobes present. Microorganism s classified as non-hydrocarbon utilizers m a y also playan im portant role in the eventual rem oval of petroleum from the environm ent. Degrad ation of petroleum involves progressive or sequential reactions, in which certain organism s m a y carry out the initial attack on the petroleum constituent; this produces interm ediate com pounds that are subsequently utilized by a different group of organism s, in the process that results in further degradation (Karrick, 1977). Biodegradation of oil com ponents As described earlier, petroleum com ponents can be classified into four m a jor groups: saturates, arom atics, resins, and asphaltenes. Major m e tabolic pathways for m a ny of these com pounds have been well studied and docum ented (Atla s, 1981 & 1984; Cerniglia, 1992; W a tkinson and Morgon, 1990) to explain their differences in susceptibility to biodegradation.
Saturates In general, the n -alkanes are the m o st readily degraded com ponents in a petroleum m i xture (Zobell, 1946; Atlas, 1981). Biodegradation of n -alkanes with m o lecular weights up to C 44 has been dem onstrated (Haines and Alexander, 1974). Alkanes in the C 10 to C 26 range are considered the m o st readily and frequently utilized hydrocarbons (Atlas 1995b;

19NAS, 1985). The predom inant m echanism of n-alkane degradation involves term inal oxidation to the corresponding alcohol, aldehydes, or fatty acid functional group. Branched alkanes are less readily degraded in com p arison to n -alkanes. Methyl branching increases the resistance to
m i crobial attack because fewer alkane degrader s can overcom e the blockage of beta-oxidation (NAS, 1985). Highly branched isoprenoid alkanes, such as pristane and phytane, which were earlier thought to be resistant to biodegradati on, have also been shown to be readily biodegradable. Cycloalkanes, however, are particularly resistant to biodegradation. Com p lex alicyclic com pounds such as hopanes and steranes are am ong the m o st persistent com pounds of petroleum spills in the environm ent (Atlas, 1981).
Arom atics Although the arom atics are generally m o re resistant to biodegradation, som e low
m o lecular weight arom atics such as naphthalene m a y actually be oxidized before m a ny saturates (Focht and W e stlake, 1987). Monoarom a tic hydro carbons are toxic to som e m i croorganism s due to their solvent action on cell m e m b ranes, but in low concentrations they are easily biodegradable under aerobic conditions. PAHs with 2-4 rings are less toxic and biodegradable at rates that decrease with the level of com p lexity. PAHs with f i ve or m o re rings can only be degraded through co-m etabolism , in which m i cr oorganism s fortuitously transform non-growth substrates while m e tabolizing sim p ler hydrocarbons or other prim ary substrates in the oil. Alkylated arom atics are degraded less rapidly than their parent com pounds; the m o re highly alkylated groups are degraded less rapidly than le ss alkylated ones. The m e tabolic pathways for the biodegradation of arom atic com pounds have b een the subject of extensive study (Atlas, 1981; Prince, 1993; Cerniglia, 1992). The bacterial de gradation of arom atics norm a lly involves the form ation of a diol, followed by ring cleavage and form ation of a di-carboxylic acid. Fungi and other eukaryotes norm a lly oxidize arom atics using m ono-oxygenases, form ing a trans-diol. Resins and asphaltenes Com p ared to saturates and arom atics, very little is known about biodegradation of resins and asphaltenes; this is due to their com p lex structures, which are difficult to analyze. Resins and asphaltenes have previously been considered to be refractory to degradation. However, there is recent evidence of asphaltene degradation through com e tabolism (Leahy and Colwell, 1990). Som e resins, particul arly low-m o lecular-weight resin fractions, can also be biodegraded at low concentrations ( NAS, 1985). Further research is still needed to understand the biodegradation of these com pounds.
In summ a ry, the susceptibility of petroleum hydro carbons to m i crobial degradation is generally in the following order: n-alkanes > branched al kanes > low-m o lecular-weight arom atics > cyclic alkanes. However, this pattern is not univers al (Perry, 1984). The com positional heterogeneity
am ong different oils greatly affects the biodegrada tion rate of their constituents. The degradation rate f o r the sam e oil constituents m a y vary signi f i cantly f o r dif f e rent oils. Com e tabolism also plays an im portant role in oil biodegradation. Many com p lex branched, cyclic, and arom atic hydrocarbons, which otherwise would not be biode graded individually, can be oxidized through
com e tabolism in an oil m i xture due to the abundan ce of other substrates that can be m e tabolized easily within the oil (Atlas, 1981). The biological fate of oil com ponents in an oil m i xture still requires further research. Particularly, effort s hould be m a de to establish a database regarding the biodegradability of different types of oils and petroleum products.

202.3.2 Environmental factors affecting oil biodegradation
Wh en oil spills occur in the environm ent, the rate of oil biodegradation is also greatly inf l uenced by the characteristics of the contam inated envir onm ent. Major environm ental factors affecting oil biodegradation include weathering processes, tem p erature, availability and concentration of nutrients, availability and concentration of oxygen, and pH. Weathering processes The weathering processes described in section 2. 2.1 have profound effects on oil biodegradation. Evaporation of volatile oil com ponents m a y be nefit m i croorganism s by rem oving m o re toxic low-m o lecular-weight com ponents such as benzene and sm aller n -alkanes. However, this process also leads to a lower biodegradable percen tage of oil, since these com ponents in general are readily biodegraded (Atlas, 1981; NAS, 1985). The oil surface area is im portant because growth of oil degraders occurs alm o st exclusiv ely at the oil-water interface (Atlas and Bartha, 1992). Form ation of water-in-oil em ulsions or m ousses reduces the surface area, therefore decreasing biodegradation. Tarballs, which are large aggregates of weathered and undegraded oil, also restrict access to m i croorganism s because of their lim ited surface area (Leahy and Colwell, 1990). Dispersion of hydrocarbons in the water columnin the form of oil-in-water
em ulsions increases the surface area of the oil and thus its availability for m i crobial attack. The form ation of oil-in-water em ulsions thr ough the m i crobial production and release of biosurfactants has also been found to be an im portant process in the uptake of hydrocarbons by bacteria and fungi (Singer and Finnerty, 1984). In contrast, the application of chem ical dispersants has produced m i xed results and has not been shown to be an effective way to enhance oil biodegradation. Photooxidation lead s to the form ation of m o re soluble com pounds, which are often m o re biodegradable. Howeve r, the effect of photooxidation processes on biodegradation is still not well understood (Nicodem et al ., 1997). Biodegradation rates are also inf l uenced by con centrations of individual oil constituents, which
m a y be affected by various weathering processes. For exam ple, m i crobes m a y attack very low concentrations of pollutants in the environm ent inefficiently (Focht and W e stlake, 1987). However, high concentrations of hydrocarbons m a y cause inhibition of biodegradation by nutrient or oxygen lim itations or toxic effects. There would seem to be, for m a ny hydrocarbons, an optim um concentration range for m e tabolism be low which degradation is not stim ulated and above which inhibition occurs. W eathering proce sses will affect the ultim ate concentrations of petroleum hydrocarbons in the environm ent in different ways. Evaporation m a y reduce the concentrations of volatile com pounds but concen trate som e other constituents. Sorption and
em ulsification m a y concentrate the pollutants, while dispersion and dissolution tend to dilute them

. Temperature The am bient tem p erature of an environm ent a f f ects both the properties of spilled oil and the activity or population of m i croorganism s. At low tem p eratures, the viscosity of the oil increases, while the volatility of toxic low-m o lecular-weight hydrocarbons is reduced, delaying the onset of

21biodegradation (Atlas, 1981). Som e hydrocarbons ar e m o re soluble at lower tem p eratures (e.g., short-chain alkanes), and som e low-m o lecular-wei ght arom atics are m o re soluble at the higher tem p erature (Focht and W e stlake, 1987). A lthough hydrocarbon biodegradation can occur over a wide range of tem p eratures, the rate of bi odegradation generally decreases with decreasing tem p erature. Highest degradation rates genera lly occur in the range of 30 to 40°C in soil environm ents, 20 to 30°C in som e freshwater environm ents, and 15 to 20°C in m a rine environm ents (Bossert and Bartha, 1984; Cooney, 1984; Jordan and Payne, 1980).
The effect of tem p erature is also com p licated by other factors such as the com position of the
m i crobial population. In environm ents where a psychrophilic population has been established, degradation can occur at significant rates unde r cold conditions. Hydrocarbon biodegradation has been observed at tem p erature as low as 0-2° C in seawater and –1.1°C in a soil. Colwell et al. (1978) reported greater degradation of Metula crude oil at 3°C than at 22°C with a m i xed culture in beach sand sam p les. W e stlake et al. (1974) also found that bacteria capable of degradation at 4°C would m e tabolize oil at 30°C, but those populations that developed at 30°C had a lim ited activity at 4°C. Oxygen Aerobic conditions are generally considered necessary for extensive degradation of oil hydrocarbons in the environm ent since m a jor de gradative pathways for both saturates and
arom atics involve oxygenases (Atlas, 1981; NAS , 1984; Cerniglia, 1992). Many studies have shown that oxygen depletion leads to sharply reduced biodegradation activities in m a rine sedim e nts and in soils (Atlas, 1981; Bossert and Bartha, 1984; Ham b rick III et al ., 1980). Conditions of oxygen lim itation norm a lly do not exist in the upper levels of the water columnin
m a rine and freshwater environm ents and in the surface layer of m o st beach environm ents. It
m a y becom e lim iting in subsurface sedim e nts, anoxic zones of water column s, and m o st finegrained m a rine shorelines, freshwater wetlands, m udflats, and salt m a rshes. Factors affecting the availability of oxygen also include the action of wave and water flow, the physical state of the oil, and the am ount of available substrates. Anaerobic oil degradation has been shown in som e studies to occur only at negligible rates, as reviewed by Atlas (1981), leading to the conc lusion that the environm ental im portance of anaerobic hydrocarbon degradation can be discounted. However, recent studies have shown that anaerobic hydrocarbon m e tabolism m a y be an im portant process in certain conditions (Head and Swannell, 1999). The biodegradation of so m e arom atic hydrocarbons, such as BTEX
com pounds, has been clearly dem onstrated to occur under a variety of anaerobic conditions (Krumholz et al ., 1996; Leathy and Colwell, 1990). Studies ha ve also dem onstrated that in som e m a rine sedim e nts, PAHs and alkanes can be degraded under sulfate-reducing conditions at sim ilar rates to those under aerobic conditions (Caldwell et al. , 1998; Coates et al ., 1997). The im portance of anaerobic biodegradation of oil in the environm ent still requires f urther studies.
Nutrients In theory, approxim a tely 150 m g of nitrogen and 30 m g of phosphorus are utilized in the conversion of 1 g of hydrocarbon to cell m a terial s (Rosenberg and Ron, 1996). Wh en a m a jor

22oil spill occurs in m a rine and freshwater envi ronm ents the supply of carbon is dram atically increased and the availability of nitrogen and phosphorus generally becom e s the lim iting factor for oil degradation (Atlas, 1984; Leahy and Colw ell, 1990). In m a rine environm ents, nutrient lim itation is generally correlated to the low background levels of nitrogen and phosphorus in seawater (Floodgate, 1984). Nutrient concentrati ons are m o re variable in freshwater system s where lakes and wetlands range from oligotrophic to eutrophic; rivers can be nutrient-poor at the source, but generally becom e nutrient-rich dow nstream after receiving industrial and dom estic effluents and agricultural runoff (Cooney, 1984). Fres hwater wetlands are typically considered to be nutrient lim ited, due to heavy dem a nd for nutrients by the plants. They are also viewed as being nutrient traps, as a substantial am ount of nutrients m a y be bound in biom ass (Mitsch and Gosselink, 1993). Both freshwater lakes and wetlands m a y also exhibit seasonal variations in nutrient levels, which will affect the perform a nce of oil biodegradation. W a rd and Brock (1976) found that in an oil-contam inated lake, oil biode gradation was at the highest rate during early spring when the nutrient content (i.e., N and P) was also high. As N and P levels decreased in the
summ e r (probably due to algal productivity) oil biodegradation also decreased. Another potential lim iting nutrient is iron, which was found to lim it oil degradation in clean offshore seawater, but is not likely to be lim iting in freshwater (Focht and W e stlake, 1987). Other factors Other im portant factors affecting biodegradation of petroleum hydrocarbons include pH and salinity. The pH of seawater is generally stable and slightly alkaline (Bossert and Bartha, 1984). In contrast, the pH of freshwater and soil envi ronm ents can vary widely. Organic soils in wetlands are of ten acidic, while m i neral soils ha ve m o re neutral and alkaline conditions. Most heterotrophic bacteria and fungi favor a neutral pH, with fungi being m o re tolerant of acidic conditions. Studies have shown that degradation of oil increases with increasing pH, and that optim um degradation occurs under slightly alka line conditions (Dibble and Bartha, 1979; Focht and W e stlake, 1987). Changes in salinity m a y affect oil biodegradati on through alteration of the m i crobial population. Dram atic variation in salinity m a y occur in es tuarine environm ents where m a rine organism s m i ngle with freshwater form s. Many freshwater organism s can survive for long periods in seawater although few can reproduce. In contrast, m o st m a rine species have an optim um salinity range of 2.5 to 3.5% and grow poor ly or not at all at salinity lower than 1.5 to 2% (Zobell, 1973). In a study of hypersaline salt evaporation ponds, W ard and Brock (1978) showed that rates of hydrocarbon m e tabolism decreased with increasing sa linity in the range of 3.3 to 28.4%. More studies are required to understand the effect of salinity on oil biodegradation.
2.3.3 Evaluation of oil biodegradation: application of biomarkers The evaluation of oil biodegradation is a difficu lt task, especially in the f i eld, due to the
com p lication of weathering processes and the heter ogeneity of contam inated sites. As described earlier, physical and chem ical weathering can significantly affect the com position and concentrations of oils. Oil contam inated sites are often highly heterogeneous, where oil concentrations can vary greatly within a sm all area. Consequently, variability associated with

23f i eld studies can be so high as to preclude or interf ere with one’s ability to discern signif i cant treatm e nt differences. Non-biodegradable or sl owly biodegradable com ponents in oil - often called biom arkers - have been used successf ully to m itigate the high variability associated with field studies (Bragg et al ., 1994; Venosa et al ., 1996; Lee et al ., 1997a). This approach estim ates the extent of biodegradation by evaluating th e ratios of target hydrocarbon concentrations relative to the concentration of these recalcitran t biom arkers. Studies have shown that use of biom arkers reduces spatial variability of oil data when com p ared to other m a ss balance approaches and allows biodegradation to be m onitored effectively by reducing the num ber of
sam p les required (Douglas et al. , 1994).
Comm only used biom arkers for evaluation of biodegradation of crude oils include the isoprenoids pristane and phytane, steranes, and th e pentacyclic triterpanes such as the hopanoids (Peters and Moldowan, 1993). Wh ile the isoprenoi ds pristane and phytane are som e what m o re resistant to biodegradation than n -alkanes with sim ilar boiling points (C 17 , C 18 ), they should only be used to m onitor the earliest stages of a bi odegradation treatm e nt program , as they are known to be biodegradable under natural conditions (Prince et al ., 1994b). Hopanes have becom e the biom arker of choice as they are m u ch m o re re sistant to m i crobial biodegradation (Atlas, 1981; Peters and Moldowan, 1993). The com pound 17 α (H), 21 β (H)-hopane was successfully used to determ ine the efficacy of biorem ediati on field trials coordinated with the Exxon Valdez spill clean up operations (Douglas et al ., 1994; Mearns, 1997; Prince, et al ., 1994 a&b). However, caution m u st be taken with their use as biom arkers since they are also very resistant to physical and chem ical weathering processes that affect m a ny alkanes and arom atics. Therefore, hopane
norm a lization is m o re usef ul in reducing the variability associated with heterogeneous oil distribution or in cases where th e effects of physical and chem ical weathering are negligible. Biodegradation can also be verified as the m a in rem oval m echanism by determ ining the relative degradation rates for hom ologous series of alkylated PAHs (Elm endorf et al ., 1994; Venosa et al. , 1997a).
2.4 Laboratory Studies on Bioremediation of Oil Biodegradation as a natural process m a y proceed slowly, depending on the type of oil (i.e., light crude oils degrade faster than heavier oils ). Biorem ediation strategies are based on the application of various m e thodologies to increase the rate or extent of the biodegradation process. The success of oil spill biorem ediation depends on our ability to optim ize various physical, chem ical, and biological conditions in the contam in ated environm ent. As described in previous sections, the m o st im portant requirem e nt is the presence of m i croorganism s with the appropriate
m e tabolic capabilities. If these m i croorganism s ar e present, then optim al rates of growth and hydrocarbon biodegradation can be sustained by en suring that adequate concentrations of nutrients and oxygen are present and that the pH is between 6 and 9 (Atlas and Bartha, 1992). The physical and chem ical characteristics of th e oil and oil surface area are also im portant determ inants of biorem ediation success. Obviously , som e of these factors can be m a nipulated
m o re easily than others. For exam ple, on an operational scale, there is nothing that can be done to alter the chem ical com position of the oil.

24There are two m a in approaches to oil spill biorem ediation. 1) bioaugmentation, in which oildegrading m i croorganism s are added to supplem ent the existing m i crobial population, and 2) biostimulation , in which the growth of indigenous oil de graders is stim ulated by the addition of nutrients or other growth-lim iting cosubstrates and/or habitat alteration. Both these approaches have been extensively studied, on a laboratory s cale as well as in the biorem ediation of oil contam inated shorelines.
2.4.1 Bioaugmentation Although hydrocarbon-degrading m i croorganism s are widespread in nature, bioaugm entation has been considered as a potential strategy for oil- biorem ediation since the 1970s. The rationale for adding oil-degrading m i croorganism s is that indigenous m i crobial populations m a y not be capable of degrading the wide range of potential substrates present in com p lex m i xtures such as petroleum (Leahy and Colwell, 1990). Other conditions under which bioaugm entation m a y be considered are when the indigenous hydrocarbon- degrading population is low, the speed of decontam ination is the prim ary factor, and whenseeding m a y reduce the lag period to start the biorem ediation process (Forsyth et al., 1995). Many vendors offer m i crobial agents claim i ng to enhance oil biodegradation (Prince, (1993). However, laboratory studies on bioaugm enta tion have produced m i xed results. Aldrett et al. (1997) tested 12 comm e rcial m i crobial cultures fo r biorem ediation of Alaska North Slope crude oil in the lab. After 28 days, f our products showed an enhancem ent of oil biodegradation with significantly higher degradation rates of alkane s and arom atics when com p ared to a nutrient control. In another shaker-flask experim e nt, Hozum i et al . (2000) investigated the effectiveness of a m i crobial product in treating a heavy oil spilled from Nakhodka using the thin layer chrom a tography-flam e ionization detection (TLC-FID) analysis

. They found that approxim a tely 35% of the oil was degraded with addition of th e m i crobial product com p ared to no oil loss for a control during a three-week test period. Surprisi ngly, the asphaltene fraction showed the highest loss am ong the four m a jor oil com ponents, which raises the question whether this oil loss was actually due to biodegradation rather than so m e quality control problem with the chem ical analysis. Som e laboratory studi es found that m i crobial seeding m a y enhance oil degradation in seawater but not in freshwater environm ents (Leahy and Colwell, 1990). To exam ine whether
m i crobial products can com p ete withthe indigenous populations, Venosa et al. (1991) tested 10 different comm e rcial m i crobial products using w eathered Alaskan crude oil in shaker flask
m i crocosm s . Although two products showed enhancem ent com p ared to a nutrient control, better degradation was observed in every case when the comm e rcial products were first sterilized, suggesting that indigenous Alaskan m i croorgani sm s were prim arily responsible for the oil biodegradation and seeded m i croorganism s seem ed to com p ete poorly with the indigenous population in the closed flask environm ent. Thus, bioaugm entation m a y be effective in benchscale studies where environm ental conditions are well controlled, but this will not guarantee its effectiveness in the field. Creation of a “superbug” that com b ines the genetic inform ation from m a ny organism s and the ability to degrade a variety of different types of hydrocarbons has also been considered. Friello

25et al . (1976) successfully produced a m u ltiplasm id-containing Pseudomonas strain capable of oxidizing aliphatic, arom atic, terpenic, and polyarom a tic hydrocarbons. Thibault and Elliot (1980) also developed a m u ltiplasm id P. putida strain that can sim u ltaneously degrade som e lighter alkanes and arom atics. However, the surviv al of such a strain in the environm ent could is questionable. More im portantly, the issues of safety, containm ent, and the potential for ecological dam a ge m u st be fully resolved befo re field testing of these organism s can be conducted (Leahy and Colwell, 1990). There is al so the problem of public perception over the release of “foreign” and especially “gene tically engineered” m i croorganism s into the environm ent.
2.4.2 Biostimulation
Biostim ulation involves the addition of rate-lim iting nutrients to accelerate the biodegradation process. In m o st shoreline ecosystem s that ha ve been heavily contam inated with hydrocarbons, nutrients are likely the lim iting factors in oil biodegradation. The m a in purpose of bench-scale treatability studies is to determ ine the type, concentration, and f r equency of addition of
am endm ents needed for m a xim u m stim ulation in the field (Venosa, 1998). Most laboratory experim e nts have shown that addition of growth lim iting nutrients, nam e ly nitrogen and phosphorus, has enhanced the rate of oil biodegradation. However, the optim al nutrient types and concentrations vary wi dely depending on the oil properties and the environm ental conditions. Wr enn et al. (1994) studied the effects of different form s of nitrogen on biodegradation of light Arabian crude oil inrespirom eters. They found that in poorly buffered seawater, nitrate is a better nitroge n source than amm onia because acid production associated with amm onia m e tabolism m a y inhibit oil biodegradation. Wh en the culture pH was controlled, the perform a nce of oil biodegrada tion was sim ilar for both am endm ents with a shorter lag tim e for amm onia addition. Ram s tad and Sveum (1995) also com p ared the effect of nitrate, amm onia, and an organic nitrogen- containing nutrient on bi odegradation of topped Statfjord crude oil in a continuous-flow seawater co lumnsystem

. W ith no c ontrol of pH in this study, nitrate was found to have the m o st pronounced effect in stim ulating oil degradation when using pristane as a biom arker. However, in a m i crocosm study, Jacks on and Pardue (1999) found that addition of amm onia appeared to be m o re eff ective than nitrate in stim ulating degradation of crude oil in salt m a rsh soils. On a weight basis, the am ount of amm onia required to achieve the
sam e increase in biodegradation as nitrate was onl y about 20%. This was attributed to the fact that amm onia is less likely to be lost from th e system by washout due to its higher adsorptive capacity to organic m a tter. Oil biodegradation largely takes place at the inte rface between oil and water. Therefore, the ef

f ectiveness of biostim ulation depends on the nutrient concentration in the interstitial pore water of oily sedim e nts (Atlas and Bartha, 1992; Bragg et al ., 1994). The nutrient concentration should be m a intained at a level high enough to facilitate bacterial growth. However, caution
m

u st be exercised as excessive concentrations of nutrients, such as amm onia, induce toxic responses in m a ny m a rine species (Pritchard et al ., 1991). Using nitrate as a biostim ulation agent, Venosa et al. (1994) determ ined that approxim a tely 1.5 to 2.0 m g N/L supported near

26m a xim a l biodegradation of heptadecane imm obilized onto sand particles in a m i crocosm study. Du et al. (1999) investigated the optim al nitrogen con centration for the biodegradation of Alaska North Slope crude oil in continuous flow beach m i crocosm s at a loading of 5g-oil/kg sand. The results showed that nitrate concentrations belo w approxim a tely 10 m gN/L lim ited the rate of oil biodegradation. The higher nutrient requirem e nt was attributed to the m o re com p lex substrate (crude oil). Ahn (1999) further studied the eff ect of nitrate concentrations under tidal flow conditions using the sam e m i crocosm s , oil, and oil loading as Du et al. (1999). Nitrate concentrations ranging from 6.25 to 400 m g N/L we re supplied sem i -diurnally to sim u late tide flow. The results from both oil analysis ( hopane as a biom arker) and m i crobial growth (phospholipid analysis) showed that the optim al nitrate concentration fed under these conditions was approxim a tely 25 m gN/L. However, the data s howed that at the end of the test (after one
m onth), an approxim a tely 80% degradation of a ll target alkanes and PAHs was achieved in all test cases that covered a range of nitrate concentr ations in the test solutions added. This result suggested that nitrate concentration m i ght only affect the rates and not the extent of oil degradation. Further research in this regard is required to optim ize biorem ediation strategies.
One of the m a in challenges associated with bios tim ulation in oil-contam inated coastal areas is
m a intaining optim al nutrient concentrations in c ontact with the oil. Oil f r om of f

s hore spills usually contam inates the intertidal zone, where the washout rate for water-soluble nutrients can be very high, and this can adversely im pact the effectiveness of biostim ulation. Many attem p ts have been m a de in the design of nutrient deliv ery system s that overcom e the washout problem s characteristic of intertidal environm ents (Pri nce, 1993). These include oleophilic and slowrelease fertilizer form ulations, as well as syst em s that rely on the subsurface flow of water through the beach (W ise et al ., 1994). Several papers have com p ared the effectiveness of these nutrient products to stim ulate hydrocarbon biodegradation rates. Croft et al. (1995) tested the ef

f i ciency of an oleophilic organic (Inipol EAP 22) and a slow-release inorganic f e rtilizer (Max Bac) and found that the oleophilic fertilizer wa s m u ch m o re effective at stim ulating oil degradation on sand than the slow-release product. Sveum and Ram s tad (1995) also found that organic products such as fish m eal and stick water (a fish m eal by-product) were m o re effective than a slow-release f e rtilizer (Mac Bac). The f a ilure f o r this slow-release f e rtilizer was attributed to the nutrient release rate being too slow to aff ect oil biodegradation. However, in som e other studies, application of organic f e rtilizers such as Inipol EAP22 also f a iled to stim ulate oil biodegradation (Sveum and Ladousse, 1989; Saffe rm an, 1991). Safferm an (1991) investigated the rates of nutrient release from several slow -release products and found that Inipol EAP22 was rapidly washed out from oiled cobble before becom i ng available to hydrocarbon-degrading bacteria. However, no attem p t was m a de in m o st of these studies to estim ate the steady-state nutrient concentrations that would result in an inte rtidal environm ent. The variable results f r om the laboratory studies indicate that the perform a nce of these products greatly depends on the nutrient release rates and the prevailing environm ental conditions. Although m u ch research has been carried out on th e biorem ediation of oil-contam inated m a rine shorelines, few studies have been conducted on oil biorem ediation in wetland environm ents. Purandare et al. (1999) conducted the only reported m i crocosm study on biostim ulation in f r eshwater wetland. They investigated dif f e rent inorganic m i neral nutrients f o r their ability to

27enhance biodegradation of a crude oil. Aquaria of 10-gallon capacity, f illed with wetland soil and planted with species of em ergent wetland pl ants, were used to sim u late natural wetlands. Two levels of water coverage were studied: (1) water level even with soil surface, and (2) water level 10 cm above the soil surface. Six treatm e nt s were evaluated for each water level: unoiled, no-nutrient control; oiled + no nutrient control; oiled + nitrate addition; oiled + nitrate + phosphate addition; oiled + amm onia addition; and oiled + amm onia + phosphate addition. Approxim a tely 14g of weathered light Arabian crude oil was added to each columnto form about a 2mm oil layer. The results show ed that for both flooded and unflooded wetland conditions, the addition of nitrate and phosphate s eem ed to enhance the degradation of oil above the natural attenuation rate. The highest biodegr adation rates of alkanes and PAHs occurred in the high water level m i crocosm s receiving n itrate and phosphorus, in which a 90% alkane reduction and 50% PAH rem oval was observed, com p ared to only 50% alkane reduction and 40% PAH rem oval in the control m i crocosm s
. Hi gher degradation of alkanes and PAHs in the high water level relative to that in the low water level suggested an increased hydrocarbon bioavailability. However, a m i crocosm study conducted by Garcia-Blanco et al . (2001a) in a sim u lated tidal salt m a rsh found that addition of nutrients did not stim ulate the biodegradation of a No. 2 fuel oil. Low oxygen availability was suggested to be the lim iting factor for oil degradation in salt m a rshes.
In summ a ry, laboratory studies have shown that biostim ulation and, in som e cases bioaugm entation, can enhance the rates of oil biodegradation, particularly in m a rine environm ents. Oxygen m a y becom e a lim iting factor in oil biodegradation under certain circum stances, such as salt m a shes and f r eshwater wetlands. However, these conclusions still need to be verified through field evaluations.
2.5 Demonstrations of Oil Bioremediation Under Field Conditions Field studies can provide the m o st convincing dem onstration of the effectiveness of oil biorem ediation since laboratory studies are not al ways able to account for num erous real world conditions such as spatial heterogeneity, biological interactions, and m a ss transf er lim itations. Com p ared to laboratory investigations, relatively f e w tests have been carried out to evaluate the effectiveness of oil biorem ediation in the fi eld because such trials are both difficult and expensive to conduct. Swannell et al. (1996) conducted the m o st extensive review available on field evaluations of oil biorem ediation in m a rine environm ents. Venosa (1998) presented an indepth critical review em phasizing problem s in the de sign and control of the existing field tests. A review by Lee (2000) addressed the potential si gnificance of enhancing plant growth (i.e. phytorem ediation) and oil m i neral aggregate fo rm ation as biostim ulation treatm e nts. Other reviews are also available (Prince, 1993; Leahy and Colwell, 1990). This section will summ a rize the latest findings from recent field studies on m arine shorelines and freshwater wetlands, as well
as m a jor points identif ied in the previous reviews.

282.5.1 Mesocosm studies
Mesocosm s or pilot-scale system s can help to si m u late actual conditions at relatively low cost, and are frequently used as bridges between m i crocosm s and field system s. Mesocosm s have been used to evaluate the effectiveness of num erous biorem ediation strategies. Bioaugmentation Unlike results from bench-scale tests, num erous m e socosm studies have dem onstrated the ineffectiveness of bioaugm entation treatm e nts. For exam ple, Tagger et al . (1983) overlaid two
m e socosm s with crude oil. One was inoculat ed with an acclim ated culture, while only indigenous populations were used in the other. Five m onths after inoculation, no statistically significant change in oil com position occurre d between the two treatm e nts. Neralla et al. (1995) investigated the effect of seeding in salt m a rsh conditions. The greenhouse experim e nt was conducted with 19-L buckets filled with m a rsh sedim e nts and actively growing Spartina alterniflora

. There were 10 duplicated treatm e nts in a total of 20 m e socosm s. Results showed that the addition of bioaugm entation products di d not enhance the degradation of weathered Arabian Lube crude oil. The conclusion was not surprising since the soil used in the experim e nt cam e from m i crocosm s used in a sim ilar study, and a large population of hydrocarbon-degrading
m i croorganism s was already present in the sedim e nt. These results again indicate that oil biodegradation is unlikely lim ited by availability of hydrocarbon degraders and that seeded
m i croorganism s could not com p ete with indigent populations.
Biostimulation Contrary to the m e socosm bioaugm entation studies, the addition of nutrients has proven to be an ef

f ective strategy f o r oil biorem ediation. Basseres et al. (1993) conducted a m e socosm trial on biorem ediation of light Arabian crude oil us ing two 600-liter tanks filled with sandy beach
m a terials at a site near the Mediterranean. An im al m eal containing 60% protein was added at 10 % w/w to one tank. A second was le ft untreated as a control. Over the 60-day period of the test, 40% of the aliphatic f r action in the treated oil was degraded whereas only 25% was degraded in the control. The num ber of hydrocarbon degrader s was found to be higher in the m eal-treated
m e socosm than in the control. However, no replicate treatm e nts and nutrient m easurem ents were perform e d during the study, so it was not po ssible to determ ine if the observed differences were statistically signif i cant.
Mendelssohn et al. (1995) conducted a greenhouse m e socosm study to determ ine the effect of oil biorem ediation products on salt m a rsh ecosystem s. A random ized block design was used with three treatm e nts (fertilizer, m i crobial products and c ontrol) at two levels of oil dosage. Each treatm e nt com b ination was replicated five tim es for a total of 30 sods of m a rsh. The results dem onstrated that the addition of a fertilizer product significantly increased the growth response of a salt m a rsh grass ( Spartina alterniflora) and the rate of soil respiration, while the m i crobial products did not significantly affect either of these processes, suggesting the biorem ediation products had neither toxic nor stim ulatory effects on the salt m a rsh environm ents.

29Purandare (1999) conducted a m e socosm study fo llowing the m i crocosm study described earlier (Purandare et al ., 1999) to further test the effectiveness of biorem ediation of an oil-contam inated freshwater wetland. Outdoor m e socosm s each m easuring 6 m x 4.5 m , filled with wetland sedim e nts, and planted with three species of em ergent wetland plants were used in the study to investigate the effect of different inorganic m i neral nutrients on biodegradation of a Louisiana crude oil. A total of 12 of thes e m e socosm s were used (3 replicates of 4 treatm e nts). The four treatm e nt strategies included a no-nutrient c ontrol, phosphate addition, nitrate + phosphate addition, and amm onia + phosphate addition. Bi odegradation rates were com puted fromhopanenorm a lized analyte data. The results showed that addition of nutrients seem ed to enhance oil biodegradation initially. However, beyond 12 weeks, the untreated control achieved a
com p arable degree of biodegradation. No conclu sion as to which nutrients were actually lim iting the biodegradation process was reached due to th e high variability in the data. The study also found that addition of nutrients led to better plant and root growth, which suggested that, although biostim ulation m a y not significantly enhan ce oil degradation in freshwater wetlands, it
m a y encourage a faster recovery of the ecosystem . It also suggested that the wetland plants m a y have out-com peted the oil-degrading m i croorganism s for nutrients and m a y have used substrates other than oil hydrocarbons for their growth (soil organic m a tter).
2.5.2 Field Demonstrations
Biorem ediation field studies are reviewed here , first with respect to bioaugm entation and biostim ulation, then by a m o re detailed discussion of four case studies that cover a wide range of shoreline types. Bioaugmentation The effectiveness of seeding has been studied in only a few field trials. Venosa et al . (1992) conducted a field test in Pr ince W illiam Sound following the Exxon Valdez spill to investigate the effectiveness of two comm e rcial m i crobial pr oducts vis-à-vis natural attenuation and nutrient addition alone. These products were selected based on a previous laboratory study (Venosa et al. , 1991). This field trial failed to dem onstrate enhanced oil biodegradation by these products. No biostim ulation occurred in the nutrient c ontrol plots either. There were no significant differences between the treatm e nt and control plot s during the 27-day trial period. However, the site where the project took place (Disk Island) was characterized as having highly weathered (degraded) oil and very calm waters, so di ssolved oxygen m a y have been lim iting, thus precluding effective biodegradation by any m eans. Other studies (Lee and Levy, 1987; Tagger et al ., 1983) suggested that exogenous m i crobial inocula are not able to com p ete successfully with indigenous populations. One approach in
overcom ing this com p etition has been proposed by Rosenberg et al . (1992). They developed a product that com b ined a polym erized ureaform a ld ehyde (F-1) with a selected oil-degrading culture capable of using this f e rtilizer as a nitr ogen source. Thus, the bacteria had a selective advantage over the indigenous population unable to utilize F-1 as nutrient source. A field trial conducted at an Israeli beach showed that this approach seem ed to be successful in enhancing oil

30biodegradation. However, conclusions were conf ounded by the lack of adequate controls in the study (Swannell et al ., 1996; Venosa, 1998).
Studies com p aring the perform a nce of bioaugm en tation and biostim ulation have suggested that nutrient addition alone had a greater effect on oil biodegradation than did the addition of
m i crobial products (Lee et al ., 1997a; Venosa et al. , 1996). ). This is probably because the
m i crobial population is rarely a lim iting factor as com p ared to the nutrients since the size of the hydrocarbon-degrading bacterial population usuallyincreases rapidly in response to oil contam ination. Lee et al . (1997a) conducted a 129-day field trial to com p are the effect of four treatm e nts on biodegradation of weathered Venture Condensate in a sandy beach. The four treatm e nts included: inorganic nutrient and oil, a m i crobial product and oil, inorganic nutrient, a
m i crobial product and oil, and oiled control. C2 -chrysenes was used as a biom arker due to the low concentration of hopane in the condensate. The results showed that periodic addition of inorganic nutrients was the m o st effective strategy for enhancing oil degradation and reducing sedim e nt toxicity, and that the full potential of the m i crobial product was lim ited by nutrient availability. A sim ilar study conducted in a wetland by Sim on et al . (1999) also show that addition of bioaugm entation agents did not enhan ce biodegradation of an Arabian m e dium crude oil. However, nutrient addition did not dem onstr ate any significant effect in their study either, suggesting other factors, such as oxygen, were lim iting oil degradation. Several other possible reasons for the failure of i nocula in degrading contam inants in nature were
summ a rized by Goldstein et al . (1985), which include: (1) the c oncentration of the contam inant
m a y be too low to support the growth of the inoc ulated species, (2) the natural environm ent m a y contain substances inhibiting growth or activity of the inocula, (3) the growth rate of the inoculated species m a y be lim ited by predation su ch as protozoa, (4) the added species m a y use other substrates in nature rather than th e targeted contam inants, and (5) the seeded
m i croorganism s m a y be unable to m ove through the pores of the sedim e nt to the contam inants.
A few field trials did claim success in dem onstr ating the effectiveness of oil bioaugm entation, such as using Alpha BioSea TM to treat the Angolan Palanca crude oil spilled from Mega Borg off Texas coast (Mauro and Wynne, 1990) and using Terra-Zym e TM in enhancing biodegradation of a heavy oil spilled f r om Nakhodka in Japan (Tsutsum i et al ., 2000). However, the success of these studies was based on either visual observation (i.e. the Mega Borg study) or digital photographic im age analysis (i.e., the Nakhodka study). No com p rehensive m onitoring program was used to verify the oil was indeed rem oved through enhanced biodegradation. The two products basically contains the sam e form ula of bacteria cultures and nutrients (Hozum i et al ., 2000). The observed visual effects m a y have been due to physical or chem ical processes such as surfactant action associated with the products (Swannell et al ., 1996).
It seem s that in m o st environm ents, indigenous oil-degrading m i croorganism s are m o re than sufficient to carry out oil biodegradation if nutrient levels and other adverse environm ental conditions do not lim it them

. Future res earch on oil bioaugm entation should focus on investigating which ecosystem s m a y be deficien t in oil degrading m i croorganism s and what types of oils or im portant oil com ponents indigenous bacteria m a y be incapable of degrading.

31Biostimulation Although laboratory and pilot-scale studies have shown that biostim ulation is a prom ising approach in enhancing oil biodegradation, the e ffectiveness of various types of nutrients and delivery strategies still require f i eld dem onstration.
Sveum and Ladousse (1989) investigated the perform a nce of Inipol EAP 22 in different types of sedim e nts. The results showed that the oleoph ilic f e rtilizer enhanced oil biodegradation in coarse-grained sedim e nts but not in fine-grained sedim e nts.
Researchers from Fisheries and Oceans-Canada(Lee and Levy, 1987; Lee and Levy, 1989; Lee and Levy, 1991; Lee and Trem bley, 1993; Lee et al ., 1995a; and Lee et al ., 1997a) conducted a series of f

i eld tests to investigate the ef f ect of dif f e rent types of f

e rtilizer and dif f e rent deliver strategies in a low energy, sandy beach or in a salt m a rsh. Their studies dem onstrated that biostim ulation using periodic addition of inorgani c f e rtilizers (amm onium nitrate and triple super phosphate) increased the rate of oil rem oval fr om beaches as m easured by changes in oil
com position relative to conserved biom arkers such as C 2 -chrysenes and/or the decline in the n C 17 /pristane and n -C 18 /phytane ratios. In contrast, the addition of the oleophilic fertilizer Inipol EAP 22 did not enhance oil degradation ina sandy beach (Lee and Levy, 1987 and 1989). Another study involved periodic addition of water-s oluble fertilizer granules (amm onium nitrate and triple super phosphate) in an attem p t to e nhance biodegradation of waxy crude oil in a lowenergy, sandy beach and in a salt m a rsh (Lee and Levy, 1991). Two concentrations of the NH 4 NO 3 were tested (0.34 and 1.36 g/L sedim e nt). The oil used was Terra Nova crude at two different levels (0.3 and 3.0%). Results from the sandy beach showed that at the lower level of oil contam ination, no enhancem ent by f e rtilizer was achieved. However, at the higher oil contam ination level, substantial oil degradation occurred in the f e rtilized plots com p ared to the unfertilized ones. Results in the salt m a rsh we re the exact opposite. Enhancem ent by fertilizer was significant at the 0.3% contam ination leve l, but no enhancem ent occurred at the 3% oil contam ination, which was attributed to the penetration of oil into the anaerobic zone where little degradation is expected. Studies on the utility and efficacy of various slow-release fertilizer form ulations also were evaluated (Lee et al ., 1993). They dem onstrated that the effects of environm ental factors controlling nutrient deliver y from the various form ulations under review (e.g., sulfur-coated urea) were the key to bior em ediation success. Another field study conducted by Lee et al. (1995a) com p ared the perform a nce of inor ganic nutrients with organic fish bonem eal f e rtilizer. These results showed that th e organic f e rtilizer had the greatest ef f ect on m i crobial growth and activity, while the inorganic nutrients were m u ch m o re effective in crude oil degradation. All these results suggest that the success of biorem ediation is case specific, depending on oil properties, the nature of the biorem ediation produc ts and the characteristics of the contam inated environm ents. Fortunately, recent studies have shown that the oil biodegradation rate depends on the nutrient concentrations in the pore water of the sedim e nts, which could provide im portant guidance for nutrient applications (Bragg et al ., 1994, Venosa, 1996). This finding m a y also explain why som e earlier trials have failed to dem onstrate the effectiveness of nutrient

32application since the nutrient concentrations in the interstitial pore water had not been m onitored and controlled in m o st of these studies. Venosa et al. (1996) found that m a intenance of a threshold nitrogen concentration of 1 - 2 m g N/L in the interstitial pore water would result in close to m a xim u m hydrocarbon biodegradation in a sandy beach (this will be discussed in detail in the section on case studies). Future resear ch on biostim ulation probably should focus on the determ ination of the optim al interstitial nutrient con centration and the best strategies to m a intain this concentration for various environm ents whenever the degradation is lim ited by nutrient availability.
In addition to the dem onstration of the efficacy of oil degradation, it is also necessary to dem onstrate that biorem ediation does not produ ce any undesired environm ental and ecological effects. There have been concerns that enhanced m i crobial degradation of oil m i ght produce toxic m e tabolic by-products. To address this concern, Lee et al . (1995b) conducted a field study using dif f e rent f e rtilizers to investigate the ef f ect of biorem ediation on the toxicity of Venture condensate stranded on sandy beach sedim e nts. Th e toxicity of the sedim e nts was m onitored using the Microtox  Solid-Phase Test. The results indicated that sedim e nt toxicity was not signif i cantly af f ected by the addition of an inor ganic f e rtilizer (amm oniumnitrate and triplesuperphosphate). However, they did observe a sl owing of the decrease in toxicity when the organic fertilizer (fishbone m eal) was applied repeatedly, which was attributed to rapid biodegradation of the fertilizer and the production of amm onia that exceeded toxicity threshold lim its.
Case studies Four representative field studies are described in m o re detail here. Three types of m a rine shorelines and one freshwater wetland are covered.
Exxon Valdez Following the grounding of the supertanker Exxon Valdez on Bligh Reef in Prince W illiam Sound in 1989, U.S. EPA, in conjunction with th e Exxon Corporation and the state of Alaska,
em barked on the largest oil spill biorem ediation proj ect ever attem p ted in the f i eld. Extensive field trials at various sites were conducted, wh ich have been well docum ented in the literature (Bragg et al. , 1994; Prince, 1993; Prince et al ., 1994; Pritchard and Costa, 1991; Swannell et al. , 1996; Venosa, 1998). Im portant findings and lessons learned from these studies are summ a rized as f o llows.

• Seeding of bioaugm entation products failed to de m onstrate enhanced oil biodegradation. It was found that oil biodegradation on the shoreline of the Prince W illiam s Sound was lim ited by the concentration of nutrients, particul arly nitrogen, and not by the absence of hydrocarbon-degrading m i croorganism s (Pritchard and Costa, 1991; Venosa et al ., 1992).

• Three types of nutrients or f e rtilizers were te sted in the f i eld: a water-soluble inorganic f e rtilizer (23:2 N:P garden f e rtilizer f o rm ul ation), a slow-release inorganic f e rtilizer (Custom b len), and an oleophilic f e rtilizer (Inipol EAP 22). Each was shown to be variably

33effective. Inipol EAP22 and Custom blen we re chosen as biorem ediation agents, and approxim a tely 50,000 kg of nitrogen and 5,000 kg of phosphorus were applied over 120 km of the oil contam inated shorelines duri ng 1989 and 1992. W ithin two weeks after the fertilizer application, the area of cobble beachtreated with Inipol EAP22 and Custom blen appeared to be visibly cleaner than the untr eated area (Pritchard and Costa, 1991; Pritchard et al ., 1992). However, it was later found that the oil coating the surfaces of the cobble had been lifted and re-deposited within th e interstices in the beach surface.

• Heterogeneous oil distribution on the contam inat ed beaches m a de it difficult to determ ine the rates of oil biodegradation using establishe d m e thods. The changes in the ratios of a hydrocarbon com ponent to a conserved internal standard or biom arker were used as the basis for determ ining the oil degradation rate. They also found that previously traditional biom arkers such as pristane and phytane degraded rapidly in Alaskan beaches. This observation rendered use of such biom arkers inef f ective in perm itting f i rm conclusions on biorem ediation effectiveness. 17 α(Η) ,21 β(Η) -hopane, the pentacyclic triterpane containing 30 carbon atom s, was used instead. Using hopane as the biom arker, Bragg et al . (1994) showed that fertilizer application accelerated the rate of oil rem oval by a factor of approxim a tely five-fold com p ared to natural attenuation. This observation was m a de, however, from sam p les repeatedly collected fr om only one site, so the statistical basis supporting this conclusion is tenuous.

• Oil biodegradation rate appeared to be depe ndent on the nitrogen concentration in the pore water of the intertidal sedim e nts, suggesting that on-site m onitoring of nutrients in the sedim e nt pore waters could provide practical guidance for nutrient applications (Bragg et al ., 1994).

• According to the EPA/Exxon/State of Alaska joint m onitoring program , biorem ediation was an environm entally sound rem e diation technique. This was based on the results of testing the toxicity of nearshore water to sensitive m a ri ne species such as Mysid shrim p , analyzing
amm onium and nitrate concentrations, eval uating the potential of algal growth, and
m onitoring oil release into nearshore water af ter the application of f

e rtilizers (Prince et al , 1994).

• The results of the f e rtilizer application f o llowing the Exxon Valdez spill generally dem onstrate that biorem ediation m a y enhance oil biodegradation on certain m a rine shorelines. However, conclusions on the e ffectiveness of biorem ediation in the Exxon Valdez study are som e what questionable, in part because the experim e ntal design was not entirely based on sound statistical principles (V enosa, 1998). Major flaws included the lack of

replication and the attem p t to determ ine too m a ny factors in a lim ited num ber of tests, resulting in the confounding of different effects. The lessons learned from the Exxon Valdez project led to the replacem ent of “post Exxon Valdez excitem e nt” with m o re scientificallyvalid approaches (Mearns, 1997).

34Delaware Field Study
The m a in purposes of this field study were to obt ain credible statistical evidence in determ ining if biorem ediation with inorganic m i neral nutrients and/or m i crobial inoculation enhances the
rem oval of crude oil, to com pute the rate at wh ich such enhancem ent takes place, and to establish engineering guidelines on how to biorem ediate an oil-contam inated sandy shoreline.
The trial was conducted on a m e dium - to coarse -grained sandy beach (environm ental sensitivity index = 4) at Fowler Beach, Delaware (located m i dway between Dover and Rehoboth Beach). A random ized com p lete block design was used in the study. Twenty 4 x 9 m plots were set up in five replicate blocks. Each block contained four treatm e nts in random order, which included: an unoiled control plot, a no-nutrient control plot (natural attenuation), a plot receiving watersoluble nutrients, and a plot receiving water-s oluble nutrients supplem ented with a natural
m i crobial inoculum from the site. W eathered Ni gerian Bonny Light crude oil was intentionally released onto 15 plots. Nutrients (NaNO 3 and Na 5 P 3 O 10 ) were applied daily through a sprinkler system at a rate designed to achieve a target of 1.5 m gN/L average interstitial pore water concentrations. Once a week, 30 L of a susp ended m i xed population of hydrocarbon-degrading bacteria was also added to the inoculum plots. Sand sam p les were collected every 14 days from the 15 oiled plots for oil analysis, and all an alytes were norm a lized to hopane. Nitrate concentrations in the interstitial pore water of oiled plots were m easured each day. Figure 2.3a shows the hopane-norm a lized concentrations of total target alkanes (i.e., the sum of all alkane analytes from n -C 10 to n -C 35, plus pristane and phytane), while Figure 2.3b shows the total target arom atics (i.e., the sum of all groups of PAHs and sulfur heterocyclics analyzable by GC/MS and their alkyl-substituted hom ologues) in the nutrient-treated, inoculum -treated, and control plots, all as a function of tim e. Although substantial hydrocarbon biodegradation occurred in the untreated plots, statistically si gnificant differences between treated and untreated plots were observed in the biodegradation rates of total alkanes and total arom atic hydrocarbons. The results also show that bioaugm entation, evenwith indigenous organism s, does not stim ulate further degradation of hydrocarbons beyond sim p le nutrient addition. The studies further dem onstrated that m a intenance of a threshold nitrogen concentration of 1 - 2 m g N/L in the interstitial pore water would perm it close to m a xim u m hydrocarbon biodegradation.
Another im portant conclusion from this study is that background nutrient concentrations at the contam inated site should be a determ ining factor in the decision to apply biorem ediation. The background nitrogen concentration at Fowler Beach was high enough to perm it close to
m a xim u m hydrocarbon biodegradation without the need to apply additional fertilizer despite the enhancem ent observed from nutrient addition. The enhanced effect, although statistically significant, was not substantial enough to have warranted a decision to im plem ent biorem ediation on a f u ll-scale basis had there been a real spill at this site. This dem onstrates that assisted biorem ediation m i ght not always be necessa ry if sufficient nutrients are naturally present at a spill site in high enough concentrations to perform natural cleanup. For coastlines having low natural input levels of nutrients, biorem ediation would be appropriate as an alternative cleanup option.

350 1 0 0 3 0 0 5 0 0 7 0 8 0 0 1 0 0 0 2 4 6 8 10 12 14 16

n g A R O M A T I C S / n g h o p a n e (b) 0 50 10 0 15 0 20 0

n g A L K A N E S / n g h o p a n e co n t r o l

nu t r ie n t s

i n o c ul um (a )

Figure 2.3. rst-order declines in (a) total target alkanes and (b) total target arom atics 2 4 6 9

Results of the Delaware field study: fi

36Biorem ediation of a Fine Sedim e nt in England The previous two cases have dem onstrated that biorem ediation can be effective for cobble and sandy shorelines, although the extent of the enhan cem ent in effectiveness is determ ined by the natural presence of nutrients. However, m u ch le ss attention has been given to fine sedim e nt m a rine shorelines such as m udfla ts. A recent study conducted by Swannell et al. (1999a) was intended to fill this gap. One of the objectives of this study was to investigate the potential of biorem ediation in treating buried oil.
The field site was located within the Stert Flats on the southwest coast of England. Twelve plots were set up on an 80-m stretch of sand com posed of 3.2 % m ud and 80% particles in the range of 125-180 μ m
. The experim e ntal area was divided in to three replicate blocks. Each block consisted of four random ly assigned treatm e nts th at consisted of an unoiled control plot, an unoiled plot treated with f e rtilizer alone, an oile d plot with no am endm ents added, and an oiled plot treated with fertilizer. The sedim e nt in each plot was retained in m e sh enclosures (0.4 m x 0.4 m x 0.05m ) and buried at a depth of 15 cm

. W eathered and em ulsified Arabian Light crude oil was applied to the appropriate enclosures at 3.7 kg/m 2 . Inorganic f e rtilizer (NaNO 3 and KH 2 PO 4 ) was applied using a sprinkling device at rate s of 2% of N and 0.2% of P by weight of the initial oil concentration, and at a f r equency of once a week f o r the f i rst f our weeks and then every two weeks thereafter. Sam p les were taken on Day 7, 49, and 108 for oil analysis, and hopane was used as a biom arker.
The results showed that m o re oil was degraded in 2 of the 3 f e rtilized plots than in the controls after 108 days. In blocks 1 and 3, the m ean total GC resolvable hydrocarbons (TGCRH)/hopane ratio decreased by 58.4% and 48.4% respectively, com p ared to 23.0% and 4.4% in the two oiled controls respectively. In block 2, oil degr adation was slower, with only 14% decrease of TGCRH/hopane ratios in the treated plot vs. no rem oval in the control plot. Statistical analysis dem onstrated that differences in the ratio of TGCRH and TPH against hopane between the fertilized plots and the controls were highly si gnificant (p<0.0001). Microbiological analyses also showed that nutrient addition increased the num bers of hydrocarbon-degraders on the oiled plots by ten

- fold. This study suggested that biorem ediation by nutrient enrichm e nt for the treatm e nt of buried oil in f i ne sedim e nts m a y be f easible af ter an oil spill incident. However, because of the third replicate failing to confirm the results of the other two, m o re definitive conclusions cannot be m a de from this study. One deficiency of this study was that by aerati ng the subsurface sedim e nts during the sedim e nt burying and periodic excavation of the site due to nutrient application, the potential problem of oxygen lim itation in this environm ent could not be evaluated. Oxygen lim itation is a m a jor concern for application of oil biorem ediation in subsurface sedim e nts, anoxic zones of a water column , and m o st fine-grained m a rine shorelines (Head and Swannell, 1999). Because this experim e ntal site was a high-energy beach w ith a tidal bore, oxygen m a y not have been a lim iting factor for this specific fine sedim e nt beach. Conclusions from this study, therefore, should only be confined to the St ert Flats and not extrapolated to sim ilar environm ents. Further

37research is still required to determ ine the ef f ectiveness of biorem ediation in other types of f

i ne sedim e nts.
Biorem ediation of a Freshwater W e tland in Canada A field study was conducted by the U.S. Envir onm ental Protection Agency, University of Cincinnati, and Fisheries and Oceans Canada at a freshwater wetland site situated along the St. Lawrence River, Ste. Croix, Quebec, Canada. The objective of this study was to determ ine the effectiveness of biostim ulation strategies in accelerating restoration of an oil-contam inated freshwater wetland site. (Garcia-Blanco et al. , 2001b, Venosa et al ., 2002). Strategies that were evaluated were biorem ediation by nutrient enrichm e nt in the presence and absence of vegetative growth of the dom inant plant species, Scirpus pungens . Twenty 5m x 4m plots were set up in the upper intertidal zone of a study site located al ong the St. Lawrence River, where the water was far enough away from the Atlantic to be still fresh, yet was tidally influenced. These plots were divided into four replicate blocks. Each bloc k consisted of five random ly assigned treatm e nts, which were: (1) oiled, no added nutrients, with intact plants (natural attenuation control), (2) oiled, NH 4 NO 3 + Ca(H 2 PO 4 ) 2 . H 2 O added, with all vegetative gr owth cut back to ground surface daily to suppress plant growth, (3) oiled, NH 4 NO 3 + Ca(H 2 PO 4 ) 2 . H 2 O added, with intact Scirpus pungens , (4) oiled, NaNO 3 + Ca(H 2 PO 4 ) 2 . H 2 O added, with intact Scirpus pungens , and (5) unoiled, NH 4 NO 3 + Ca(H 2 PO 4 ) 2 . H 2 O added, with intact Scirpus pungens (background control). W eathered Mesa light crude oil was released ont o each plot earm a rked for oiling. The am ount of oil released was 12 L per plot. Com posite core sam p le s were collected after 0, 1, 2, 4, 6, 8, 12, 16, and 21 weeks f o r quantif ication of the rem a ining oil constituents by gas chrom a tography/m a ss spectroscopy (GC/MS) operating in the Selected Ion Monitoring m ode or SIM. To account for differences due to physical washout, all oil constituents were norm a lized to hopane. Figure 2.4 illustrates results from the hopane-norm a lized concentrations of total target alkanes and total target PAHs for the four treatm e nts as a function of tim e. Although the biorem ediation and phytorem ediation treatm e nts achieved sligh tly better degradation of hydrocarbons than natural attenuation, no statistically significan t evidence of stim ulation through addition of nutrients or biodegradation enhancem ent by vege tation was observed. After 21 weeks, reduction of target parent and alkyl-substituted polycyclic arom atic hydrocarbons (PAHs) averaged 32% in all treatm e nts. Reduction of total target alkane s was of a sim ilar m a gnitude. The pattern of disappearance of hydrocarbons was characteristic of biodegradation; nam e ly, the lower
m o lecular weight alkanes declined to a greater extent than the higher carbon-num ber alkanes, as did the lower m o lecular weight PAHs com p ared to the higher m o lecular weight PAHs. Since there was little evidence to support enhancem ent of biodegradation by nutrient addition with and without vegetation, it was suggested that oxygen lim itation was m o st likely the dom inant cause of the persistence of oil hydrocarbons on the oil-contam inated plots.

38

### 600

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0 0 1 2 4 6 8 12 16 21 N a tu r a l A t te n u a tio n Bi o + Am m o ni u m Ph yto + A m m o ni u m Ph yto + N i t r ate (a)

(b ) 

### Nat u r a l At t e n u a t i o n Bi o + Am m o ni um P h yt o + A m m o nium Phyt o + Ni t r at e



### Tim e (w eek s) Figure 2.4. Results of St. Lawren ce River field study: concentration declines in (a) total target alkanes and (b) total target PAHsngPAHs/nghopanengAlkanes/nghopane

39Wh ile the Ste. Croix results indicate that biostim ulation m i ght not be an effective strategy to
m e diate the rem oval of residual oil from the sedim e nts, significant changes in biological
m easures of habitat were observed (Lee et al ., 2001a). For exam ple, S. pungens , the dom inant plant species, was tolerant to the oil, and its gr owth was significantly enhanced above that of the unoiled control by the addition of nutrien ts. Other biotest organism s (bacteria, Vibrio sp ., invertebrates, Daphnia, Hyalella, and Viviparus sp .) provided additional evidence of both enhanced recovery and potential detrim ental eff ects. The study indicates that on an operational scale, natural attenuation m a y be the m o st practical treatm e nt option for oxygen-lim ited freshwater wetlands.
2.5.3 Kinetics of oil bioremediation Knowledge of the kinetics of oil biodegradation under different environm ental conditions is im portant for assessing the potential fate of targeted com pounds, evaluating the efficacy of biorem ediation, and determ ining appropriate strategies to enhance oil biodegradation. Oil biodegradation rates are difficult to predict due to the com p lexity of the environm ent. The rates of biodegradation vary greatly am ong the vari ous com ponents of crude oils and petroleum products. The presence of other substrates m a y a ffect the degradation rates of the com pounds of interest. Environm ental factors such as tem p er ature, nutrient concentrations, and oxygen tension also influence the kinetics of oil degradation. The heterogeneity of oil distribution on shorelines or wetland sedim e nts m a kes kinetics studies even m o re difficult. Very few kinetic studies on oil degradation unde r field conditions have been conducted. The Exxon Valdez m onitoring program developed a m u ltiple regression m odel based on f i eld studies conducted by researchers from Exxon (Prince et al ., 1994). The best-fitting m odel was expressed as:
C h (t) = α [1-p(t)] γ e δ r ( t )+ω t ε ( 2.2)
where C h (t) is the tim e-varying hopane-norm a lized con centration of an analyte, p is the polar f r action of the oil, r is the ratio of the average residual nitrogen concentration to oil loading, and ε is the assum e d m u ltiplicative error term , while α , δ , γ , and ω are f itting param e ters determ ined f r om the m u ltiple regression analysis. This m odel m a tched the experim e ntal results in Alaska well when the param e ters are chosen to f it the data. However, its potential f o r process understanding and prediction is lim ited because the data set used for the regression was lim ited to only one sm all, non-replicated area in the field.
Venosa et al. (1996) developed from fiel d data first-order biodegradation rate constants for resolvable alkanes and im portant two- and three -ring PAH groups present in light crude oil. The first order relationship was expressed as:  A   A   e − kt ( 2

. 3 )   =   H   H  0
40where (A/H) is the tim e-varying hopane-norm a liz ed concentration of an analyte, (A/H) 0 is that quantity at tim e zero, and k is the first-order biodegradation rate constant for an analyte. For the field study conducted in Delaware, first- order biodegradation rate coefficients ranged from 0.026 to 0.056 day -1 for total resolvable alkanes and from 0.021 to 0.031 day -1 f o r total resolvable PAHs (Venosa et al ., 1996).
Actual 1 st -order biodegradation rates are not constant , however. Instead, theyare a function of the residual nutrient concentration:  N  k obs = k max    K n + N    ( 2

. 4 )

where k obs and k ma x (T -1 ) are the observed and m a xim um fi rst-order hydrocarbon biodegradation rates, respectively, K n(M n L -3 ) is the half-saturation concentrat ion for a specific nutrient, and N (M n L -3 ) is the interstitial pore water residual nutrient concentration. Experim e nts conducted at the University of Cincinnati showed that the K nfor nitrate is approxim a tely 0.5 m g N/L (unpublished). The m odel that incorporates Equa tion 2.3 and 2.4 will be very useful in the experim e ntal design and the perform ance predicti on for the biorem ediation of oil contam inated shorelines.
Studies have also been conducted to com p are oil biodegradation rates obtained in laboratory tests to those calculated from the Delaware field study (Venosa et al ., 1996, 1997a; Holder et al ., 1999). Venosa et al . (1996, 1997a) found that the degradati on rates of all target alkanes and
arom atics in a light crude oil were close to an or der of m a gnitude lower in the field com p ared to results from the laboratory. However, when the ra te data of PAHs were norm a lized to the highest alkyl-substituted hom ologue in each given PAH series, the first order rate constants in the field were nearly identical to rate constants f r om th e lab. These relationships were consistent even with different m i crobial consortia isolated from eight different m a rine shorelines of the United States. Sim ilar results were also observed in a laboratory study using 14 different m arine and freshwater consortia (Holder et al ., 1999).
Sim on et al . (1999) reported som e kinetic data derived from a study conducted in a coastal wetland contam inated with Arabian light crude oil. First-order biodegradation rate coefficients ranging from 0.017 to 0.061 day -1 f o r total target saturates and f r om 0.009 to 0.027 day -1 f o r total target arom atics were reported. These rate coefficients were sim ilar to those of Venosa et al

. (1996). Further research is still required to devel op m o re state-of-art m odels and to establish a database of kinetic param e ters for different types of oil under various m arine shoreline and freshwater wetland environm ents.
2.6 Nutrient Hydrodynamics
Since nutrient addition has been found to be the m o st effective biorem ediation strategy in aerobic environm ents, particularly for m arine shorelines , a full understanding of the fate of water soluble

41nutrients on m arine beaches and the hydrodynam i cs controlling their transport and persistence is necessary. One of the m a in challenges associat ed with biostim ulation in oil-contam inated coastal areas is m a intaining optim um nutrient c oncentrations in contact with the oil and the degrading m i croorganism s. Oil from offshore spil ls usually contam inates the upper third of the intertidal zone, where the washout rate for wa ter-soluble nutrients can be very high. Various oleophilic and slow-release nutrient f o rm ulations have been developed to im prove the contact between oil and nutrients within the environm ent. However, m o st slow-release and m a ny oleophilic f e rtilizers rely on dissolution of the nutri ents into the aqueous phase bef o re they can be used by hydrocarbon degraders (Safferm an , 1991). Thus, design of effective oil biorem ediation strategies and nutrient delive ry system s requires an understanding of the transport of water-soluble fertilizers in a beach ecosystem

. Dissolved nutrients are expected to m ove with the water in the beach sand. W a ter flow through the porous m a trix of a beach is driven by a com b ination of three m a in factors (Boufadel et al. , 1999b; Wr enn et al ., 1997): (1) tides that result in rise and fall of beach groundwater level (typically, the water level in a beach tracks the leve l of the rising tide with only a slight lag, but the beach drains m u ch m o re slowly when the tid e ebbs because of resistance from the porous
m a trix), (2) wave action that operates thr ough two m a in m echanism s (at wave runup, water enters the beach and percolates vertically thr ough the unsaturated zone until it reaches the water table; at wave rundown, water m oves in a predom inantly horizontal direction and exits at the water line), and (3) flow of fr esh groundwater from coastal aquifers, which causes continuous horizontal advective flow from the beach face at or near the water line (Glover, 1959). This type of groundwater flow can interact with tidal fluc tuations to produce com p lex variations in the groundwater level within the beach (See Section 2.6.1).
2.6.1 Nutrient transport in beaches: a mesocosm study Beach hydraulics and nutrient hydrodynam i cs were investigated through both theoretical and experim e ntal approaches by Boufadel et al . (1999 a and b, 1998, 1997). A two-dim e nsional f i nite elem ent m odel f o r water f l ow and salt tran sport in saturated and unsaturated porous m e dia was developed. The m odel also considers the e ffects of salt concentration on water density and water viscosity. An experim e ntal wave tank was used to validate the m odel and to investigate cases that cannot be sim u lated by the num erical m odel. Tracer tests were carried out to investigate the separate and com b ined effects of tide, waves, and buoyancy on the transport of soluble inorganic nutrients in sand beaches. The m a jor findings from these studies were:

• In the absence of regional seaward groundwater flow, the tide generated a predom inantly downward and seaward hydraulic gradient that caused the washout to the sea of nutrients applied to the top section of beaches. The presence of waves under these conditions accelerated the washout, the rate of which was found to increase by about 30% when waves were superim posed on the tide.

• Beach geom etry plays a m a jor role in beachhydraulics and hydrodynam i cs because the flow lines are perpendicular to the beach surface. Under tidal action, seawater enters the beach

42from above, flowing vertically downward, and m a y cause the entrapm e nt of less saline water in the beach. This finding is very im portant because system s that rely on continuous or interm ittent injection of nutrient solutions thr ough trenches or horizontal wells installed in the supratidal zone rely on subsurface flow to carry the nutrients through the contam inated area (W ise et al ., 1994). The approach that was proposed by W i se et al . (1994) assum e s that nutrients dissolved in the freshwater plum e will be brought into contact with the oiled beach
m a terial periodically by the rising tide, because the freshwater plum e should float on top of the saltwater. The finding of freshwater trappe d between two saltwater wedges indicates that subsurface injection of nutrients m ay not be an effective m e thod for providing nutrients to the biorem ediation zone when a freshwater plum exists because the im pact zone will never be exposed to the nutrients.

• Different nutrient application strategies at low tide were investigated in the m e socosm beach since application of nutrients at the beach surface at low tide (Venosa et al. , 1996) appears to be the best application strategy for biostim u lati on. The results showed that applying nutrients at the beach surface pre-dissolved in water resulted in generally longer residence tim es and larger spreading of the nutrient plum e in the top section of the beach com p ared to applying nutrients in granular form at the beach surface and hosing them in with a water spray. However, it should be noted that in practice, add ition of granular fertilizer m ay be prudent in cases where it is not possible to sprinkle pre-di ssolved nutrients as the application m e thod.
2.6.2 Nutrient transport in beaches: field trials To verify the findings from the m e socosm studies and to further investigate nutrient transport under field conditions, tracer studies were conducted in the intertidal zone of three different
m arine beaches in Delaware and Maine (Suidan and Wr enn; 2001; Wr enn et al .,1997a & b). The Delaware Tracer Study The purpose of this study was to characterize th e transport of water-soluble nutrients in the intertidal region and to estim ate their washout rates from the biorem ediation zone (i.e., the oilcontam inated area). The study was conducted on a m oderate-energy, sandy beach (Slaughter Beach) on Delaware Bay. The typical wave heights at this beach were between 15 to 30 cm . A conservative tracer (LiNO 3 ) was applied to eight replicat e 5 m x 10 m plots in the upper intertidal zone at low tide during full m oon spri ng tide and the last-quarter m oon neap tide. The tops of the plots were placed approxim a tely at the spring high tide line. Sand sam p les were collected for tracer analysis. This study showed that the rate of tracer washoutfrom the biorem ediation zone was m o re rapid when the tracer was applied at spring tide (when the tidal am plitude is largest) than at neap tide. Wh en the conservative tracer (LiNO 3 ) was applied to the beach surface in the upper intertidal zone at the full m oon spring tide, it was com p le tely rem oved within one day. Wh en it was applied at neap tide, however, the tracer persiste d in the biorem ediation zone for several days. The results indicated that the am ount of nutri ent rem a ining in the biorem ediation zone was highly correlated with the m a xim um extent to which the treated area had previously been

43subm erged by water at high tide; subm ergence resu lted in nearly com p lete rem oval of dissolved
com pounds from the biorem ediation zone. Theref ore, the fertilizer release rates should be designed to achieve optim al nutrient concentrati ons while the tide is out. The Delaware tracer study indicates that nutrient tran sport in sandy beaches is driven by tidally influenced hydraulic gradients and wave activity. However, it is im possi ble to clearly separate the influence of tide and wave action in this study. The Maine Tracer Study The m a in purpose of this field study was to separate ly evaluate the effect of tide and wave action on nutrient transport since the results of the Slaughter Beach field studies could not distinguish between these two processes. This was acco m p lished by com p aring the transport rate and characteristics of a dissolved conservative tracer ( lithium nitrate) on a high-energy beach to those on a low-energy beach to determ ine how waves a ffect solute transport (Suidan and Wr enn, 2001; Wr enn et al., 1997b). The two beaches are located in the town of Scarborough, in southern Maine. Scarborough Beach is a high-energy beach that faces the Atlantic Ocean with average heights of breaking waves between0.3 to 1 m during the course of this study, whereas Ferry Beach faces a protected harbor with the typical wa ve height of less than 3 cm . The tidal range at both beaches was essentially the sam e . Dissolved tracer (LiNO 3 ) was applied to four replicate plots on each beach at low tide during full m oon spring tide and the last-quarter m oon neap tide. Both water and sand sam p les were collected for tracer analysis.
W a shout of lithium from the upper intertidal zone during the spring-tide experim e nt is shown in Figure 2.5. The differences between the two beaches are very clear. Wh ereas lithium was com p letely rem oved from the entire experim e ntal dom ain within two days on the high-energy beach, m o re than two weeks were required to ach ieve the sam e degree of washout fromthe lowenergy beach. W a shout during the neap-tide expe rim e nt was m u ch slower than that observed during the spring-tide experim e nt. Lithium was com p letely rem oved from the experim e ntal dom ain on Scarborough Beach within one week , whereas the total m a ss of lithium was essentially unchanged for m o re than two weeks at Ferry Beach. Slower washout was expected during the neap-tide experim e nt, because only the bottom s of the plots were covered by water at high tide during m o st of the first five or six days

. Since plot coverage by the tide was essentially the sam e on the two beaches during the first week of this experim e nt, the differences in washout rate m u st have been due prim arily to wave activity. The Maine field study clearly shows that the was hout rate of nutrients from the biorem ediation zone is strongly affected by the wave activity of the contam inated beach. W a ve action in the upper intertidal zone m a y cause nutrients from th e surface layers of the beach to be diluted directly into the water column , resulting in their imm e diate loss f r om the biorem ediation zone. On the other hand, washout due to tidal activityalone is relatively slow, and nutrients will probably rem a in in contact with oiled beach m a terial long enough to effectively stim ulate oil biodegradation on low-energy beaches.

44051015time(days)0.511.52.5.54time(days)(a)FerryBeach(b)ScarboroughBeach0233(b)ScarboroughBeach(a)FerryBeachFigure 2.5 Washout of Lithiumfromthe upper intertidal zone during the spring-tide experiment at(a) Ferry Beach and (b) Scarborough BeachThe mass remaining isshown for (A) inthe totalexperimentaldomain(B) inside the plotboundaries, and (C) inthe bioremediation zone(i.e., within 30 cmof the surface).45BBBBBBEEEEBEBEBBBBBBBBBBEEEEEEEECBBBBBBBBBEEEEEEEEEBBBBBBBBBBBEEEEEEEEAEBSpring-1Spring-2EBSpring-1Spring-2EEBEEE3.52.530time(days)time(days)1.520.51150BBBEEE050010001500BE050010001500

L i t h i u m ma s s r e ma i n i n g ( g )BE05001000150001002003004005006000100200300400500600
lit h i u m m a s s r e m a in i n g ( g )0100200300400500600700BBBBBBBBBBBBBBBBEEEEEEEEEEEEEEEE050010001500CBBBBBBBBBBBBBBBBEEEEEEEEEEEEEEE050010001500

L i t h i u m ma s s r e ma i n i n g ( g )BBBBBBBBBBBBBBBBBEEEEEEEEEEEEEEEE050010001500AEBSpring-1Spring-2EBSpring-1Spring-201002003004005006000100200300400500600
lit h i u m m a s s r e m a in i n g ( g )0100200300400500600700BEBEEBE5104Salinity distribution was also exam ined in the Maine study. In both beaches, the “sandwiching phenom enon” (a layer of lower salinity water lo cated between two higher salinity water layers) was observed, which was confirm e d by Boufadel et al . (1999b) in the m e socosm study. Both these m e socosm and field studies are very he lpful in providing guidelines for the optim al applications of nutrients on m arine beaches, which will be discussed later in this docum ent. Although m any questions rem a in unanswered, we have m a de trem endous progress in understanding various aspects of oil biorem ediation in the last decade. The developm ent of an operational guideline for biorem ediation of oil cont am inated m arine shorelines and freshwater wetlands is not only possible but also necessary inensuring that full-scale cleanup in the future proceeds rapidly and efficiently.

46Chapter 3 Methods Used in Monitoring Oil Bioremediation
In order to dem onstrate that biodegradation is taking place in the field, the chem istry or
m i crobial population m u st be shown to change in ways that would be predicted if biorem ediation were occurring (NRC, 1993). Environm ental conditi ons, particularly nutrient concentrations, should also be m onitored for evaluating the e ffects of biorem ediation. Som e of the m o st im portant m e thods used in m onitoring oil biorem ediation will be overviewed in this chapter. References will be given for detailed descriptions of m e thods.
3.1 Analytical Methods
3.1.1 Microbiological analysis Existing m e thods for m i crobial analysis can be classified into cultured-based techniques and culture-independent techniques. Comm only used m i crobial analysis m e thods in m onitoring oil biorem ediation are summ arized in this section and table 3.1. 3.1.1.1 Enumeration of hydrocarbon-degrading microorganisms: culture based techniques
Microbial counts are often used to m onitor the biorem ediation process. In general, the m o re m i crobes, the m o re quickly the contam inants will be degraded. Correlating an increase in the
num ber of contam inant-degrading bacteria above norm a l field conditions is one indicator that biorem ediation is taking place. Analysis of the m i crobial comm unities that take part in in-situ hydrocarbon biodegradation activities has been a challenge to m i crobiologists (Macnaughton et al ., 1999). The reason for this is that m o st (~ 90 to 99%) of the species m a king up com p etent degrading comm unities do not form colonies when current laboratory-based culture techniques are used (Rollins and Colwell, 1986; Rozsak and Colwell, 1987; W i lkinson, 1988). The techniques are briefly described below.
Plate count Plate count is a traditional technique, which quantif ies the num ber of bacter ia capable of growing on a prescribed set of nutrients and substrates in a solid m e dium , by counting the colonies form ed (National Research Council, 1993). The ge neral procedure involves (1) m a king the solid
m e dium or gel from a liquid solution with appropriate nutrients and substrates, using a solidifying agent like agar, (2) Spreading a sam p le containing the bacteria of interest thinly over the surface of the gel in plates, (3) Incubating the plates, (4) counting the bacterial colonies form ed. Each colony is assum e d to have arisen from a single bacterial cell.
A num ber of studies have used hydrocarbon incorpor ated into either agar-based or silica-based
m e dia to enum erate of hydrocarbon-degrading m i croorganism s (Horowitz and Atlas, 1978; Sexstone and Atlas, 1977; W a lker and Colwell, 1976). However, other researchers reported that plate counts are unsuitable for enum erating hydr ocarbon-utilizing m i croorganism s because m any m arine bacteria can grow and produce m i cro- colonies on sm all am ounts of organic m a tter existing in the solid m e dia, resulting in th e counting of non-hydrocarbon utilizers (Atlas, 1981; Higashihara, et al. , 1978). Plate counts also underestim ate the num ber and diversity of bacteria because of the difficulty in enriching viable co lonies from environm ental sam p les. Culturable

47techniques have been found to be inferior to t echniques that do not rely on viable culturing for
enum eration (Macnaughton et al., 1999). Most-probable-number (MPN) procedures MPN procedures have been viewed as a m o re reliable m e thod for enum erating hydrocarbonutilizing m i croorganism s because such procedures elim inate the need for a solidifying agent and perm it direct assessm ent of th e ability to actually utilize hydrocarbons (Atlas, 1981; Wr enn and Venosa, 1996). MPN procedures use liquid nutrient m e dia in test tubes or m i crotiter plates and hydrocarbons as the sole carbon source. The enum eration is carried out through a statistical analysis based on the num bers of a series of diluted liquid sam p les that show evidence of bacterial growth. This evidence of bacterial grow th can be established based on turbidity, release of 14 CO 2 from radiolabeled hydrocarbons, disruption of oil sheen, and reduction of dyes (Rice and Hemm i ngsen, 1997). Either statistical tables (Eaton et al ., 1995) or a com puter program (Klee, 1993) can be used to determ ine the MPN. Most existing MPN procedures use crude oil or a refined petroleum product as the selected hydrocarbons, which can not distinguish diffe rent groups of hydrocarbon degraders. For
exam ple, the Sheen Screen Method uses dispersion or em ulsification of the crude oil substrate to identify positive wells (Brown and Braddock, 1990). But these effects are associated prim arily with growth on aliphatic hydrocarbons (Homm e l, 1990). Aliphatic hydrocarbons are of less environm ental concern than PAHs, however, becau se they are less toxic and are biodegraded
m o re rapidly. Wr enn and Venosa (1996) recen tly developed an MPN procedure that can separately enum erate aliphatic and arom atic hydr ocarbon-degrading bacteria. The size of the two populations is estim ated using separate 96-we ll m i crotiter plates. The alkane-degrader MPN
m e thod uses hexadecane as the selective growth substrate, and positive wells are detected by reduction of iodonitrotetrazolium violet, which is added after incubation for 2 weeks at 20 o C. PAH degraders are grown on a m i xture of PAHs in separate plates. Positive wells turn yellow to greenish brown from accum u lation of the partial oxidation products of the arom atic substrates after 3 weeks incubation. This m e thod is sim p le enough for use in the fi eld and provides reliable estim ates for the density and com position of specific hydrocarbon-degrading populations. For a detailed description of existing MPN procedures, readers can refer to Rice and Hemm i ngsen (1997), and Wr enn and Venosa (1996). 3.1.1.2 Culture-independent population/community techniques
The m a in challenge for accurate analysis of hydrocarbon-degraders using existing culture-based techniques is that m o st these species are notable to be cultured (Atlas & Bartha, 1987, Macnaughton et al ., 1999). The em ergent culture-independentm o lecular techniques have m a de it possible to identify the diversity and com position of uncultivated m i crobial comm unities and to
enum erate bacteria in m o re precise ways. Phospholipid fatty acid (PLFA) analysis Phospholipid fatty acid (PLFA) analysis is based on the characteristic “signature” of fatty acids present in the m e m b ranes of all cells (National Research Council, 1993). The distribution of fatty acids is unique and stable. Therefore, it can be us ed as an identifying index. Determ ination of

48biom ass through analysis of the extractable lipi ds avoids culture bias. This technique also provides a quantitative m eans to m easure viable biom ass, comm unity com position, and nutritional stature (White et al ., 1998). Phospholipids can be extracted from the sa m p le and the phosphate can be m e asured by colorim e tric techniques (Findlay et al ., 1989). The results can represent the am ounts of viable cells and biological activities in the sam p le. A m o re powerful PLFA m e thod involves extraction and separation of lipid classes into neutral- , glyco-, and polar-lipid fractions, followed by quantitative analysis using gas chrom a togr aphy/m a ss spectrom e try (GC/MS) (Macnaughton et al ., 1999; Wh ite et al ., 1998). This procedure can quantitatively determ ine the characteristics of
m i crobial comm unities. However, PLFA analysis cannot identify species com position. Nucleic acid-based molecular techniques Nucleic acid-based m o lecular techniques can identi fy bacterial species by the unique sequence of
m o lecular codes in their genes. One of m o st useful m e thods for determ ining the diversity of bacterial comm unities is denaturing gradient gel electrophoresis (DGGE) (Muyzer et al., 1993). The m e thod provides a m e ans of separating the P CR (polym erase chain reaction) products from
m i xed cultures based on the m e lting properties of the DNA. Usually the 16s rDNA portion of the bacterial genom e is targeted for PCR am plification, since this region is comm only used for bacterial identif ication. PCR prim ers can be de signed to detect a broad range of bacteria (universal) or can be designed just for a specific group of interest. The DGGE gel is m a de of acrylam ide, and contai ns a gradient of form am ide and urea, which both act to denature, or pull the strands of the DNA apart. The PCR products are loaded onto the gel, and a voltage is established across the gel for several hours. As DNA, which carries a net negative charge, is carried through the gel, it en counters an increasing gradient of denaturant, which causes the DNA chains to separate and the effective size of the m o lecules increases, causing m ovem e nt through the gel to cease. The end result is a DNA banding pattern, where DNA requiring m o re chem ical potential to dena ture travels farther, and DNA requiring less chem ical potential to denature stays near the top of the gel. Each DNA band approxim a tely corresponds to the presence of one kind of organism in the m i xed culture. This banding pattern is som e tim es referred to as a “comm unity fingerp rint,” and allows for a quick approxim a tion of
num ber of bacterial species (diversity) present. Bands can be excised for sequencing analysis, and sequences can be com p ared to the Ribosom al Database Project (Maidak et al., 2000) containing the 16S rDNA sequences of currently known organism s. The use of DGGE for quantitative purposes is still not well established. However, it can be used in conjunction with other quantitative m e thods suchas PLFA analysis to provide insight into
m i crobial species distribution. The PLFA- DGGE techniques were successfully used in determ ination of m i crobial population changes during the Delaware field study on oil biorem ediation (Macnaughton et al ., 1999).

49Table 3.1 Comm only used m i crobial analysis m e thods in m onitoring oil biorem ediation Method Advantages
Plate counts

Most probable num ber (MPN) techniques

Phospholipid fatty acid (PLFA) analysis

Denaturing gradient gel electrophoresis (DGGE) W e ll-established and easy to perform

More reliable m e thod for
enum erating hydrocarbonutilizing m i croorganism s since hydrocarbons are used as the sole carbon source. Som e of the procedures can separately
enum erate aliphatic and
arom atic hydrocarbondegrading bacteria
Elim inate culture bias and quantitatively determ ine viable biom ass, comm unity
com position, and nutritional stature Identify species distribution without culture bias Counts only organism s viable on solid-m edia and m ay allow growth of non-hydrocarbon degraders Relatively labor-intensive (large num ber of incubations) and tim e consum ing, and m a y still be subject to culture bias

Require specified knowledge and equipm ent, and can not identify species com position

Also require m o re specified knowledge and equipm ent, and quantitative analysis is still not well-established Disadvantages

3.1.2. Chemical analysis of nutrients Since oil biodegradation is lim ited by availabil ity of nutrients in m o st m arine shorelines,
m onitoring nutrients, particularly the nutrient c oncentrations in pore water, is critical in developing proper biorem ediation strategies and assessing the effect of oil biorem ediation (Bragg et al ., 1994; Venosa et al ., 1996). Im portant nutrient analyses include m easurem ents of
amm onium , nitrate, nitrite, and phosphorus. Comm onl y used m e thods for these nutrient analyses
are summ arized as follows. Sample preparation Analysis of nutrients in sedim e nts can either be conducted on site or be frozen and shipped to a lab for m easurem ents. Before the analysis, av ailable nitrogen and phosphorus species can be extracted from sedim e nts using a 2M solution of KCl or an acidified 0.1% NaCl solution (Page et al. , 1986; Tan, 1996). Total nitrogen and phosphorus can be liberated from sedim e nt sam p les by persulfate digestion at 121°C.

50Ammonia analysis The m o st comm only used m e thods for amm onia an alysis are autom a ted colorim e tric m e thods due to their high sensitivity and ease of use. Two m a jor autom a ted colorim e tric m e thods are available (1) Autom a ted phenate m e thod (4500-NH 3 H, Eaton et al ., 1995) (2) Autom a ted salicylate-hypochlorite m e thod. The latter was r ecently developed due to the environm ental concerns associated with the phenol used by the phenate m e thod (Tan, 1996). The extracted
amm onia nitrogen can also be m easured in the field using an amm onia-selective electrode (4500- NH 3 F, Eaton et al ., 1995) or som e comm e rcial kits, such as a Chem etrics kit (Chem e trics Inc., Calverton, VA) although these m e thods are less sensitive and m o re susceptible to interference than colorim e tric m e thods. Nitrate and nitrite analysis Comm only used techniques for nitrate analysis include the ultraviolet spectrophotom etric
m e thod (4500-NO 3 - B, Eaton et al ., 1995), autom a ted cadm i um reduction m e thod (4500-NO 3 - F,

- Eaton et al ., 1995) and nitrate electrode m e thod (4500-NO 3 D, Eaton et al ., 1995). The ultraviolet spectrophotom etric m e thod is suitable for rapid m easurem ent or screening sam p les that have low organic m a tter contents. Th e autom a ted cadm i um reduction m e thod is m o re sensitive and suitable f o r nitrate analysis in vari ous types of water and wastewater. Wh en using this m e thod, nitrate is reduced to nitrite by passing through a Cu-Cd reduction column . The nitrite is then determ ined using a colorim e tric procedure. Therefore, without the reduction step, this m e thod can be used for nitrite analysis. Tota l nitrogen can also be m e asured as nitrate by the autom a ted cadm i um reduction m e thod following oxidation of urea, amm onia, and nitrite by ° potassium persulfate at 121 C (4500-N or g D, Eaton et al ., 1995). The nitrate electrode m e thod can be used in the field for rapid nitrate analysis although it is less reliable than cadm ium reduction
m e thod. Phosphorus analysis Phosphorus analysis involves two general proce dures: (1) conversion of the phosphorus form of interest to dissolved orthophosphate and (2) colorim e tric determ ination of dissolved orthophosphate (Eaton et al ., 1995). Total phosphorus analysis usually uses persulfate digestion procedures (m ethod (4500-P B, Eaton et al ., 1995). Various digestion procedures have been developed for analysis of Available Phosphorus (Tan, 1996), a variable concept that reflects the
am ount of phosphorus available to plants or m i croorganism s. The ascorbic acid m e thod (4500-P E, Eaton et al ., 1995) is recomm e nded for analysis of phosphate in the concentration range of 0.01 to 6 m g P/L. Comm e rcial kits, such as Hach ® phosphate analysis procedures (Hach Com p any, Loveland, CO), are also available for use in the field.
3.1.3 Chemical analysis of oil and oil constituents One of the prim ary m easures of the success of biorem ediation treatm e nts is reduction in the concentrations of spilled oils and target oil cons tituents in particular. Various techniques have been developed and used in petroleum hydr ocarbon analysis, which include gravim etric
m e thods, infrared spectroscopy (IR), gas chrom a tography/flam e ionizati on detection (GC/FID), gas chrom a tography-m a ss spectrom e try (GC/MS ), and thin-layer chrom a tography-flam e ionization detection (TLC-FID). Oil analysis m e thods can be generally classified into two

51categories: nonspecific m e thods to m easure tota l petroleum hydrocarbons (TPHs), and specific
m e thods using various chrom a tographic techniques to quantify target oil constituents. Comm only used oil analysis m e thods in m onitoring oil biorem ediation are summ arized in this section and table 3.2. 3.1.3.1 Total petroleum hydrocarbon (TPH) techniques TPH techniques m a inly include gravim etric and infrared spectroscopic m e thods. These techniques are widely accepted m e thods to rapidl y quantify the oil due to their sim p licity and low costs. However, these m e thods provide little inform ation about oil com ponents, exhibit high detection lim its and are susceptible to various interferences (Douglas et al 1991, Xie et al ., 1999). Furtherm ore, they are unable to distinguish between abiotic and biotic losses.
Gravim etric analysis involves solvent (e.g., di chlorom e thane) extraction, evaporation, and gravim etric m easurem ent (EPA 413.1 and EPA 9071) . This m e thod does not distinguish between petroleum hydrocarbons and naturally occurring biogenic com pounds (such as plant lipids) and
m ay result in overestim ating TPH. Infrared spectroscopy (IR) involves solvent extraction
norm a lly using trichlorotrifluoroethane (Fre on 113). TPH is subsequently m easured by
com p aring the infrared absorption of the extrac tion liquid against that of a defined hydrocarbon
m i xture (e.g., EPA 418.1 (U.S. EPA, 1992)). Although this technique is a m o re sensitive
m easure of hydrocarbons than gravim etric m e thods, it m ay also overestim ate or underestim ate TPH for a variety of reasons (Xie et al ., 1999). Environm ental concerns regarding the use of Freon as a solvent is also a potential problem for its application (Rom ero and Ferrer, 1999). Fully halogenated solvents such as tetrachloroethene or tetrachlorom e thane can be substituted f o r Freon.
Comm e rcial kits, such as Petrof lag ® test kit (Dexsil, Ham d en, CT) and Hach ® DR/2000 test kit (Hach Com p any, Loveland, CO), are also availa ble f o r use in the f i eld with lim ited reliability (Lam bert et al ., 1999a&b). All TPH techniques are severely affected by the spatial heterogeneity. A larger quantity of oil at one s pot could give a m i sleadi ng high TPH value. They do not distinguish between abiotic and biotic losses that are im portant in correctly interpreting the data f o r the f a te of the petroleum in the envi ronm ent. So, one has to be careful in carrying out these tests as well as in interpreting the data. A sufficiently large num ber of sam p les m ay help to
overcom e som e of the variability. 3.1.3.2 Analysis of specific oil constituents
To assess the effect of oil biorem ediation, id entification and quantification of individual oil
com ponents and com pounds, particularly those const ituents that are of significant environm ental concern , is required. Various chrom a tographic t echniques, particularly GC/FID and GC/MS, are widely used to provide specific and sensiti ve analysis of oil constituents (Douglas et al. , 1994; W ang et al . 1997). Gas chromatography/flame ionization detection (GC/FID) GC/FID com b ines chrom a tographic separati on of hydrocarbon fractions on a capillary GC columnwith quantification by FID (e.g., EPA 8100(U.S. EPA, 1992)). Pretreatm e nt of oiled

52sedim e nt sam p les involves drying of sam p les, add ition of surrogates, and solvent extraction (e.g., EPA 3540c (U.S. EPA, 1992)). This m e thod has beenm a inly used for detection of aliphatic hydrocarbons such as individual C 10 to C 35 n-alkanes, and isoprenoid hydrocarbons (Table 3.3). GC/FID can also be used to determ ine TPHs by the m e thod of internal standards (Douglas et al , 1994). Total GC-detectable hydrocarbons m a ke up m o re than 50% of the oil com ponents (Prince et al ., 1994). However, GC/FID m ay not be used to identify and quantify PAHs and biom arker
com pounds because it can not clearly separate m any of these com pounds, especially the alkylated PAHs. Since PAHs are of m o st envi ronm ental concerns and biom arker com pounds are critical in distinguishing between biodegrada tion from physical weathering processes and reducing spatial variability of oil data, m easur em ent of these com pounds is very im portant in
m onitoring oil biorem ediation. Gas chromatography-mass spectrometry (GC/MS) GC/MS, which com b ines chem ical separation by GC and spectral resolution by MS, is the choice of m e thods for specific com pound determ ination, es pecially for identification and quantification of PAHs and biom arkers (W ang et al ., 1997). The m a ss spectrom e ter is often operated in the selected ion m onitoring m ode (SIM) to f urther increase sensitivity and selectivity relative to conventional full-scan GC/MS. For a detailed de scription of GC/MS analytical procedures, readers can refer to Douglas et al . (1994), EPA Method 8270 (US EPA,1989), and Venosa et al

. (1996). Typical target alkanes and PAHs anal yzed using GC/MS in recent USEPA-sponsored projects are listed in Table 3.3 and 3.4 (Purandare, 1999; Venosa et al ., 1996). Because the distribution of oil on shorelines contam inated by offshore spills can be highly heterogeneous (Bragg et al ., 1994), the concentrations of all target analytes are often reported relative to a conservative biom arker such as 17 α (H),21 β (H)-hopane (Douglas et al ., 1994). Detailed discussions on biom arkers will be presented in Section 3.2. Thin-layer chromatography flame ionization detection (TLC-FID) Thin layer chrom a tography-flam e ionization de tection (TLC-FID), which uses a special instrum e nt called the Iatroscan, separates hydr ocarbons on a Chrom arod thin layer based on characteristic chem ical types or fractions suchas aliphatic, arom atic, polar, and asphaltene
com pounds (Goto et al ., 1994). This m e thod has advantages in m easuring high-boiling-point hydrocarbons such as higher m o lecular weight saturates, arom atics, resins, and asphaltenes,
som e of which m ay not be detectable by GC or HPLC. Unlike the other analytical techniques, which are either too gross (e.g. TPH techniques) or very specific (e.g. GC based m e thods), the TLC-FID can m easure the relative percentages of th e four m a jor fractions of petroleum in a short period of tim e. This m e thod has been successfully used for m onitoring oil biorem ediation in a wetland environm ent (Stephens et al ., 1999). However, TLC-FID can not identify specific
com pounds and m ay only be used as a screening tool. Controversy also exists concerning the reliability of the m e thods due to som e confusing results, and m odifications of the technique have been suggested (Cebolla et al. , 1998).

53Table 3.2 Comm only used m e thods for petroleum hydrocarbon analysis Method Advantages Disadvantages References TPH techniques Inexpensive and easy to Low sensitivity and USEPA, 1992

GC/FID

GC/MS

TLC/FID perform

. Used as quick screen tools

A specific m e thod used for detection of aliphatic and a lim ited num ber of
arom atic hydrocarbons. Highly sensitive and selective m e thod for identification and quantification of a wide range of hydrocarbons, including PAHs and biom arkers Quick detection of a wide range of oil com ponents, including high m o lecular weight saturates, arom atics, resin, and asphaltenes; can be used as an effective screening tool. selectivity; can not be used f o r identif ication of oil com ponents. Not recomm e nded.
Unable to identif yand quantify alkylated PAHs and biom arkers

Expensive equipm ent and
com p licated procedures

Unable to identif y specific com pounds; quantitative analysis is still not well-established Douglas et al. , 1994; W ang et al .,1997

Douglas et al. , 1994; USEPA, 1992; Venosa et al. , 1996; W a ng et al .,1997

Stephens et al ., 1999; Cebolla et al. , 1998

54Table 3.3 Target alkanes list (Purandare, 1999) Compound Name Misc. Info. QIon Response Factors Reference Compounds Internal Standards
D26 n-dodecane surrogate 66 D26 n-dodecane D22 ndecane QIon: 66 n C10 57 n C10

n C11 57 n C11

n C12 57 n C12

n C13 57 n C13

n C14 57 n C14

n C15 57 n C15 D36 n-heptadecane surrogate 66 D36 n-heptadecane D34 n-hexadecane QIon: 66 n C17 57 n C17 Pristane 57 Pristane

n C18 57 n C18 Phytane 57 Phytane

n C19 57 n C19 D50 n-tetracosane surrogate 66 D50 n-tetracosane D42 neicosane QIon: 66 n C20 57 n C20

n C21 57 n C21

n C22 57 n C22

n C23 57 n C23

n C24 57 n C24

n C25 57 n C25

n C26 57 n C26

n C27 57 n C27 D66 n-dotriacontane surrogate 66 D66-dotriacontane D62 n-triacontane Qion: 66 n C28 57 n C28

n C29 57 n C29

n C30 57 n C30

n C31 57 n C31

n C32 57 n C32

n C33 57 n C33

n C34 57 n C34
n C35 57 n C35 5 –cholestane surrogate 217 5 -cholestane 5 -antrostane QIon: 245 Ho p ane alkane 191 Ho p ane

55Table 3.4 Target PAHs analyzed by GC/MS (Purandare, 1999) Compound Name Misc. Info. QIon Factors Reference Compounds Internal Standards
D10 1-m e thyl naphthalene surrogate 152 D10 1-m e thyl naphthalene Naphthalene PAH 128 Naphthalene C1 naphthalene 142 Naphthalene D8- naphthalene QIon: 136

C2 naphthalene 2-ring alkyl PAHs 156 C3 naphthalene 170 Naphthalene C4 naphthalene 184 Naphthalene Fluorene 3 ring PAH 166 Fluorene C1 fluorenes 180 Fluorene C2 fluorenes 3 ring alkyl PAHs 194 D10- anthracene QIon: 188
C3 fluorene 208 Fluorene Dibenzothiophene 3 ring P AH Dibenzothiophene C1 dibenzothiophene 198 Dibenzothiophene C2 dibenzothiophene 3 ring alkyl PAHs 212 C3 dibenzothiophene 226 Dibenzothiophene D10 phenanthrene surrogate 188 D10 phenanthrene Phenanthrene PAHs 178 Phenanthrene Anthracene 178 Anthracene C1 phenanthrenes 192 Phenanthrene C2 phenanthrenes 3 ring alkyl PAHs 206 C3 phenanthrenes 220 Phenanthrene C4 phenanthrenes 234 Phenanthrene Naphthobenzothiophene ring PAH Dibenzothiophene C1 naphthobenzothiophene 248 Dibenzothiophene C2 naphthobenzothiophene 4 ring alkyl PAHs 262 C3 naphthobenzothiophene 276 Dibenzothiophene Fluoranthene PAH 202 Fluoranthene D10 pyrene surrogate 212 D10 pyrene D12- chrysene Q Ion: 240 Pyrene PAH 202 Pyrene C1 pyrenes 216 Pyrene C2 pyrenes 4 ring alky PAHs 230 Response

2-ring
Naphthalene

Fluorene

184
Dibenzothiophene

3-ring

Phenanthrene

4 234
Dibenzothiophene
4-ring

4-ring
Pyrene

56Table 3.4. Target PAHs (Contd.) Compound Name Misc. Info. QIon Response Factors Reference Compounds Internal Standards Chrysene 4 ring PAH 228 Chrysene D12- chrysene QIon: 240 C1 chrysenes 4 ring alkyl PAHs 242 Chrysene C2 chrysenes 256 Chrysene C3 chrysenes 270 Chrysene C4 chrysenes 284 Chrysene benzo(b)fluranthene 5-ring PAHs 252 Benzo(b)fluoranthene D12- perylene QIon: 264 benzo(k)fluoranthene 252 Benzo(k)fluoranthene benzo(e)pyrene 252 Benzo(e)pyrene benzo(a)pyrene 252 Benzo(a)pyrene indeno(1, 2, 3-cd) pyrene 6-ring PAHs 276 Indeno (1, 2, 3-cd) pyrene dibenzo(a,h)anthracene 5-ring PAHs 278 Dibenzo (a,h)anthracene benzo(g,h,i)perylene 6-ring PAHs 276 Benzo(g,h,i)perylene

3.2 Biomarkers
As m e ntioned in Chapter 2, internal biom arkers have been widely used to distinguish between biodegradation and the physical or chem ical loss of oil from treated plots in biorem ediation field studies (Bragg et al ., 1994; Venosa et al ., 1996; Lee et al ., 1997b). An ideal biom arker should (1) provide source specific inform ation, (2) not be form ed during physical , chem ical weathering and biological processes, (3) be non-biodegradable or relatively resistant to biodegradation on tim e scales relevant to the study or cleanup, (4 ) be extracted from the sam p le with sim ilar efficiency to the other associated com pounds (Douglas et al , 1994; Prince et al ., 1994a,b). The ideal biom arker would also be subject to the sa m e physical and chem ical rem oval m e chanism s as the target analytes so that any differences coul d be attributed to biodegradation. Unfortunately, existing biom arkers can rarely m e et all these criteria. An overview of comm only used biom arkers in m onitoring oil biorem ediation is gi ven here. Issues to be aware of when using biom arkers will be discussed.
3.2.1 Commonly used biomarkers Pristane and phytane Hydrocarbon degrading m i croorganism s usually degrade branched alkanes or isoprenoid
com pounds such as pristane and phytane at lower rates than straight-chain alkanes. Pristane and phytane are also subject to the sam e physical and chem ical rem oval m echanism s as their

57corresponding straight-chain alkanes (W ang et al ., 1998). Therefore, n-C17:pristane and nC18:phytane ratios have been traditionally used to interpret the extent of biodegradation (Gaudlach E.R et al ., 1983; Lee and Levy, 1987,1989,1991). However, it was later found in Prince W illiam Sound that these com pounds were resistant to degradation in the initial tim e period, but they degraded rather rapidly ove r longer periods of tim e (Prince 1993, Bragg, et al ., 1994). This phenom enon was also observed in a study on biorem ediation of Light Arabian crude oil in a freshwater wetland environm ent as shown in Figure 3.1 (Purandare, 1999). The study
exam ined the dry weight norm a lized pristane, phytane , and hopane concentrations over the 32- week experim e ntal period. It can be seen th at the pristane and phytane showed significant degradation, but hopane rem a ined constant th roughout the course of the study. These results show that the n-alkane:isoprenoid ratio m ay be us eful indications of biodegradation in very short term or the earliest part of a study. However, they m ay substantially underestim ate the extent of biodegradation in a long run.

20

16

12

8

4

0 0 w eek s 2 w eek s 4 w eek s 8 w eek s 1 6 w e ek s 3 2 w eek s pr y s t a ne ph y t a n e ho pa ne

W eek s o f O p er a t i o n Figure 3.1 Change of pristane, phytane and hopane in a freshwater wetland

Hopanes Hopanes, derived from the m o lecular fossils of prokaryotic and eukaryotic m e m b ranes (Peters and Moldowan, 1993), are very resistant to biodegradation as shown in Figure 3.1. 17 α (H),21 β (H)-hopane (Figure 3.2) has been found to be neither generated nor biodegraded during the biodegradation of crude oil on tim e scales relevant to estim ating the cleaning of oil spills and theref ore has appropriate characteristics to serve as an internal standard f o r m onitoringmg/kg dry weight of sediment

58biodegradation of both specific petroleum co m pounds and total oil in crude oil in the environm ent (Prince et al ., 1994). Hopanes have been viewed as the biom arkers of choice (Mearns, 1997) since the successful application of 17 α (H),21 β (H)-hopane for evaluating oil biorem ediation af ter the Exxon Valdez spill (Douglas et al ., 1994; Prince, et al ., 1994a&b). Many recent studies have also chosen this hopane as the biom arker (Garcia-Blanco et al ., 2001b; Purandare, 1999; Swannell et al ., 1999a; Venosa et al. , 1996). Figure 3.2 shows alkane analysis results in soil core sam p les during the first 21 weeks of the St. Lawrence River field study usi ng dry weight soil norm a liza tion and hopane norm a lization (Garcia-Blanco et al ., 2001b). Due to the substantial heterogeneity of oil distribution in the freshwater wetland sedim e nts, large standard de viations (17-72% of the m ean concentrations) were obtained when using dry weight soil norm a li zation. No convincing conclusions were able to be reached based on these data. However, the st andard deviation of oil concentrations in the sedim e nt sam p les were m u ch lower (2-15% of the m ean oil concentrations) when using hopane
norm a lization, which enabled evaluation of oil bi odegradation with high levels of statistical confidence. However, it should be noted that hopanes are also very resistant to those physical and chem ical weathering processes that affect m o st targ et alkanes and arom atics (e.g., dissolution, volatilization, and photooxidation). Therefore, al though hopane norm a lization is very useful inreducing the variability associated with heterogeneous oil distribution, changes in the hopanenorm a lized analyte concentrations alone m ay not be used to verify that biodegradation is the prim ary rem oval m echanism

. Use of hopane norm a lization to distinguish biodegradation from physical loss of oil is valid m o st ly when effects of dissolution and volatilization are negligible (Prince, 1993; Venosa et al ., 1996), which is not always true

. Other m eans of verifying biodegradation are needed such as the use of alkylated PAH isom ers (see next section). Alkylated PAH isomers Biodegradation can be verified as a rem oval m echanism by determ ining the relative degradation rates for hom ologous series of alkylated PAHs. Preferential biodegradation of aliphatic and
arom atic hydrocarbons based on m o lecular structure has long been recognized (Jobson et al., 1972; W a lker et al., 1976; Roubal and Atlas, 1978; Fedorak and W e stla ke, 1981; Elm e ndorf et al., 1994, W ang et al, 1998). This is particularly true for alkylated PAHs, which are biologically transform e d m o re slowly as the extent of al kyl substitution increases (Elm endorf et al., 1994; Venosa et al., 1997). Biodegradation results in unique characteristic changes in the distributions of hom ologous series of alkylated PAHs. On th e other hand, physical w eathering does not cause the sam e types of alterations in their relative distributions (W ang et al., 1998). In other words, physical weathering causes equal losses in all hom ologues irrespective of the extent of alkyl substitutions. Recent research has also shown that the relative biodegradation rates for alkylated hom ologs of the 2- and 3-ring PAHs were rem a rkably sim ilar for m i xed cultures of hydrocarbondegrading bacteria isolated from a wide variet y of sources (Venosa et al., 1997), and they were also very sim ilar in the f i eld, despite the m u chhigher absolute rates that were observed in the laboratory (Venosa et al., 1996). Therefore, the distribution patterns of alkylated PAHs, when used in conjunction with other oil analysis data, canbe very useful in accurate assessm ent of the extent and progress of oil biodegradation.

59a) D r y s o i l w e ight norm al i zed

1 000

800

600

400

200

0 0 1 2 4 6 8 1 2 1 6 2 1 N a tu ral A tten u a tio n Bi o + Am m oni um P hyt o + Am m oni um P hyt o + Ni t r a t e Unoi l e d + Am m oni um

Ti me ( w e e k s ) b) H o pa n e n o r m a l i z e d

60 0

50 0

40 0

30 0

20 0

10 0

0 0 1 2 4 6 8 12 16 21 N a tu ral A t ten u atio n B i o + Am m oni um Phyt o + Am m oni um Phyt o + Ni t r a t e

Ti m e ( w e e k s ) Figure 3.2 Com p arison of alkanes analysis results in soil core sam p les from St. Lawrence field study using (a) dry soil weight norm a lization and (b) hopane norm a lizationmgalkanes/ mghopanemgalkanes / Kgdry soil

60Other Biomarkers For refined petroleum products such as diesel fu el and fuel oil #2 that do not contain the hopane, C 4 -phenanthrenes/anthracenes m ay be substituted (Douglas et al ., 1994). These com pounds are degraded, but very slowly. Lee et al . (1997b) used C2-chrysenes as biom arkers in their studies on biorem ediation of weathered Venture Condensat e because hopane concentration in the oil was at or near the detection lim its of their instrum e nts and C2-chrysenes are also considered highly degradation resistant. C4-chrysenes can also be used if their concentrations are high enough to be detectable.
Table 3.5 Biom arkers and their characteristics Biom arkers Lim itations Applicable to References Pristane & Phytane

Hopanes

Alkylated PAHs isom ers

Phenanthrenes, Anthracenes, Resistant biodegradation

Loss characteristic of biodegradation Biodegradable

Can distinguish biodegradation from abiotic weathering processes Monitoring early stages of biodegradation

Reducing spatial variability

Verification of biodegradation

Refined petroleum Lee et 1987&1989

Prince et , 1994; et al., 1994

Venosa et al ., 1997a; W ang et al ., 1997&1998

Douglas et al ., 1994 Advantages

to

pattern not al.

al. Douglas

and Chrysenes products

3.2.2 The effect of contaminant redistribution on observed remediation rates Loss of oil due to physical washout and sand redistribution can be significant in an oiled beach. Hopane norm a lization is an effective way to dis tinguish biodegradation from the effects of the physical washout and sedim e nt exchange between the inside and outside of experim e ntal plots when all of the oil is ini tially present inside of the plots and m o st of the beach is clean, such as in a study involving intentional oiling (Venosa et al ., 1996). However, it will not work well in a study involving sm all plots set up on a beach contam inated by a “spill of opportunity” or a real oil spill. The reason is that when a study is carri ed out on a beach com p letely contam inated by an oil spill, oil or oiled sand will transport between relatively sm all treated areas (i.e., experim e ntal plots) and large untreated but oiled areas. Si nce hopane is a conservative biom arker, its concentration will be the sam e inside and outside of the plots (assum i ng that the beaches were

61uniform ly oiled initially). Therefore, oiled sand com i ng into the plots will have the sam e am ount of hopane as the sand leaving the plots (assum i ng that treatm e nt does not result in physical
rem oval of bulk oil f r om sand inside the plots), butthe concentrations of target analytes will be higher, because the biodegradation rates will be lower in untreated areas. Since the hopane concentration inside the plots will not be af f ect ed by sand transport, it does not allow us to quantify the rate of sand exchange between the inside and outside of the plots. A theoretical analysis conducted by Wr enn et al . (1999) illustrates how physical exchange of sand between treated and untreated areas of the be ach affects the observed biodegradation rate of target analytes when sm all plots are set up on a large beach. For the nutrient-treated plots, the rate of change in the analyte concentrations, A treat , is given by:
dA treat A treat dT  1 dH 1 dT 

dt = T dt − k treat A treat +   H dt − T dt   A treat ( 3

. 1 )

For the untreated control plots and in the untreat ed areas of the beach, the rate of analyte, A con , disappearance is: dA con A con  dH 

dt = H   dt   − k con A con ( 3

. 2 )

Wh ere k treat and k conare the first-order biodegradation ra te coefficients in the presence and absence of nutrients; H is the concentration of hopane; and T is the concentration of a hypothetical nonbiodegradable tracer that is present in the oil inside of the plots but not outside of the plots (e.g., a hydrophobic fluorescent dye that is added to the oiled sand inside the plots at the start of treatm e nt). Equation (3.1) describes the rate of change of the analyte concentration inside the treated plots due to transport of treated oiled sand out of the plots, biodegradation, and transport of untreated oiled sand into the plots. Assum i ng the rates of physical loss of treated oiled sand from inside the plots and loss of oiled sand from the beach to be first-order processes, the hopane-norm a lized analyte concentration inside of the plots at any tim e can be solved as:  A treat   A o  ( k

− k T − k treat ) t  ( k treat − k con ) + ( k T − k B ) e ( k T + k treat − k con − k B ) t    H   =    H o    e   k T + k treat − k con − k B   ( 3

. 3 )
Wh ere k T and kH are the first-order loss coefficients for the nonbiodegradable tracer and hopane, respectively. Representative results f r om this m odel are show n in Figure 3.3 using param e ters obtained from Delaware f i eld trial (Venosa et al. , 1996). It can be seen that the effect of exchange of oiled sand between the inside of treated plots and untreated beach is to reduce the observed degradation rate relative to the true rate. Sand exchange has no e ffect on the observed biodegradation rate in the control plots, because biodegradation is assum e d to occur at the sam e rates inside and outside of those plots. The apparent rate of rem e diation in the treated plots, however, will decrease while relatively large am ounts of oil rem a in even whenbiorem ediation would be capable of achieving
a com p lete cleanup if the entire contam inated s horeline were treated. This could lead to the

62incorrect conclusion that, whereas biorem ediati on can stim ulate the initial cleanup rate, it cannot restore the contam inated shorelines to accepta ble conditions. The analysis shows that, if behavior of this type is observed, it is probably an artif act of treating a sm all f r action of the total contam inated area, and m o re com p lete rem e diation would be expected when a larger area is treated. Nevertheless, there are other possible e xplanations for incom p lete rem e diation. These include the inability to m a intain suf f i cient nutrients in the biorem ediation zone and, especially on high-energy beaches, a high loss rate of bacteria from the oiled surfaces (e.g., due to scouring by waves). Therefore, a thorough m onitoring program is a very im portant com ponent of this research, because data on nutrient concentrations and m i crobial activity is required to properly interpret the results if the cleanup goals are not achieved during the field study.

50

40

30

20

10

0 0 1 0 2 0 3 0 4 0 5 0 6 0 7 0 8 0 9 0 1 0 0 true de gradation rate ( treate d ) observed de gradation rate (treated ) contro l deg r a d ation rate

(tru e and observed )

Ti m e (d ays) Figure 3.3 Reduction in the observed oil biodegrad ation rate due to exchange of oiled sand between the inside of treated plots and untreated surrounding beach
3.3 Sampling in the Field
Because oil contam inated sites are highly heter ogeneous, representative sam p ling is difficult but also extrem ely im portant for proper evaluation of biorem ediation. Field sam p ling procedures
m

u st be designed to achieve statistically va lid sam p ling and to m i nim i ze contam ination or changes in the sam p les. Variables that affect the representativeness of sam p les and their m e thods of collection include characteristics of m e dia, c oncentration distribution of analytes, and bias introduced during collection, preparation , and transportation (Lee, 2000). General principles to achieving statistically valid sam p ling in soil environm ents and solid waste have been welldocum ented (Tan, 1996; USEPA, 1992). However, little inform ation is available regarding
sam p ling protocols for m onitoring oil biorem ediation in m arine shorelines and freshwaterTotal PAHs(ng/nghopane)

63wetlands. Major considerations to achieve repr esentative sam p ling are summ arized as follows based on the soil-science literature and field experiences from recent oil biorem ediation projects.
3.3.1 Sampling strategies Types of sampling Sam p ling m e thods can be classified into sim p le random sam p ling, system atic sam p ling, stratified
sam p ling, and com positing (Tan, 1996). The Si m p le random sam p ling or the grab m e thod involves collecting sam p les at random in a sam p ling area. This m e thod depends com p letely on the luck of the draw without considering the variation of analytes in the sam p ling field. Purandare (1999) found that results of oil analysis in a wetland sedim e ntwere highly variable and unreliable using this sam p ling approach during the early stage of his study. Therefore, this approach m ay only be suitable for use in rela tively hom ogeneous system s. System atic sam p ling involves taking sam p les based on certain patterns, such as collecting sam p les in a grid pattern. This m e thod will ensure that the entire sam p ling area is represented in the sam p le. Stratified
sam p ling involves dividing the sam p ling field into a num ber of sectors or quadrants and taking independent sam p les in each sector according to the rule of proportionality (e.g., taking m o re sam p les in m o re heavily oiled sites). These appr oaches can often provide m o re accurate results than sim p le random sam p ling, because with this m e thod the sam p les are distributed m o re evenly over the population. Com positing involves the m i xing of sam p ling units to form a single sam p le, which has the advantage of increased accuracythrough the use of large num bers of sam p ling units per sam p le. This approach in com b ination of stratified sam p ling has been frequently used in recent field studies on oil biorem ediation (V enosa et al., 1996; Garcia-Blanco et al., 2001b). Depth of sampling Sam p ling depth in oil spill sites m a inly depends on the distribution of the analytes of interest, especially the depth of oil penetration. Crude oil rarely penetrates coastal sedim e nts to depths of greater than one foot (Gundlach, 1987). Penetrati on of oil in wetland environm ents will be even less deep than in m o st m arine sedim e nts. Purandare (1999) found that oil only penetrated to 2.5
cm in a wetland sedim e nt in 16 weeks. The top 2 cm layers of the sam p le cores were then used for oil analysis. Therefore, a surv ey of oil penetration in the contam inated site is critical in determ ination of sam p ling procedures for a biorem ediation application. Size of sampling The size of the sam p le required depends on the available resources, the required degree of confidence, and the objectives of the analys is (Rupp and Jones, 1993). Generally, the m o re heterogeneous the system , the m o re intense m u st be the sam p ling efforts to reach a given accuracy. However, econom ic considerations ofte n restrict both the quantity and the num ber of
sam p les taken, and a balance should be obtained between the size of sam p les to be taken for required confidence and econom ic factors. Fo llowing expressions are som e exam ples of statistical approaches, which can be used to calculate the required num ber of sam p les with respect to an acceptable error (Peterson and Calvin, 1986; Tan, 1996):

n = 4 σ /E 2 ( 3

. 4 )

or n = t 2 S 2 /E 2 (3.5)

64where n is num ber of sam p les, t is t- test value, E is acceptable error, S 2 is sum of squares of
sam p le deviation, and σ is standard deviation. Sample handling and storage Sam p le handling procedures m u st be designed to m i nim i ze cross-contam ination or chem ical and biological changes in the sam p les. For exam ple, different sets of sam p ling tools should be used for different treatm e nts to avoid cross contam ina tion. Clean and proper containers, such as PVC bags and tin cans, should be used for sam p le storag e. Sam p les, if not being analyzed in the field, should be frozen using blue ice or dry ice until they reach their destination. And once they arrive the destination, sam p les should be stored in a freezer at -18 to - 20 o C until needed.
3.3.2 Field sampling experiences
Two exam ples of well-designed sam p ling protoc ols used in definitive biorem ediation field studies are shown below (Venosa et al. , 1996; Garcia-Blanco et al ., 2001b). Sampling protocols used in the Delaware field bioremediation study (Sandy Beach) Each plot was divided into 4 quadrants fo r sam p ling purposes, and the sand sam p les were collected at the nodes of a 0.5 m by 0.5 m grid in each quadrant (Figure 3.4). As Figure 3.4 shows, all sam p le nodes were at least 0.5 m from the plot boundary on all sides. This buffer zone was designed to m i nim i ze the im pact of edge effects on the observed extent of biodegradation. Since the plots were at least 9.5 m long, this provided a m i nim um of 28 sam p le nodes for each quadrant.
Sam p les were com posited from 2 random ly sele cted nodes in each quadrant at 1 or 2 week intervals for a period of 14 weeks, but they were never collected from the sam e node twice. The
sam p ling frequency was higher near the beginning of treatm ent and followed the order of W eeks 0 (i.e., just before treatm e nt began), 1, 2, 3, 4, 6, 8, 10, 12, and 14. W i th 10 sam p le events in this study, 20 sam p le nodes were required per quadrant. Sand sam p les were collected with hand augers to a depth of 15 cm , which was determ ined based on prelim inary oil penetration study. Sampling protocols used in St. Lawrence River field study (Freshwater Wetland) A random ized sam p ling plan was designed to elim inate sam p ling bias. Each plot was divided into six sectors, each m easuring 1.5 m x 1.0 m , with the 1.5 m dim e nsion parallel to the shoreline. Each sector was subdivided into 10 subsam pling zones, corresponding to the predeterm i ned sam p ling events. Each subsam pling zone had dim e nsions 50 cm x 30 cm

. At each sam p ling event, a 9-cm core sam p le wa s collected using a tulip bulb planter frompreassigned random subsam pling zones from each of the six sectors within each plot. These sam p les
were com b ined into 2 com posites (3 predeterm i ned sam p les per com posite). Both com posite
sam p les were placed into quart size paint cans, frozen, and shipped to the University of Cincinnati (OH) for oil analysis. Sam p les were kept in the freezer (-18 to -20 o C) until they were extracted. Com posite 1 served as the sam p le to be analyzed by GC/MS. Com posite 2 was analyzed for three of the sam p ling events to check within-plot variability. The rem a ining
com posite 2 sam p les were frozen and archived. The first sam p ling event (week 0) was carried

65out at low tide the day af ter the application of oil and nutrients. Subsequent sam p ling events were at weeks 1, 2, 4, 6, 8, 12, 16, 21, 48, and 65.
The sam p ling designs m e ntioned above are im portant because they incorporate replicate plots with random placem ent on the experim e ntal plan e. These are essential for perm itting proper statistical analysis of treatm e nt effects.

Figure 3.4 Exam ple of an experim e ntal plot in the Delaware field study showing sam p ling quadrants and nodes

3.4 Monitoring General Site Background Conditions Monitoring general site background conditions is very im portant for properly evaluating effects of oil biorem ediation. Major background conditi ons include dissolved oxygen, pore water pH, tem p erature, and salinity.

663.4.1 Oxygen Oxygen availability is crucial for rapid bior em ediation because hydrocarbon biodegradation is prim arily an aerobic process. Therefore, th e dissolved oxygen (DO) of pore water should be
m onitored on a regular basis. W a ter sam p les from the oil-contam inated region of the subsurface can be collected through the m u lti-port sam p le wells and sealed in DO bottles. Conventional
m e thods for DO m easurem ent include iodom etric procedures and m e m b rane electrode m e thod (Eaton et al ., 1995).
The iodom etric technique is the m o st precise and reliable titrim e tric procedure for DO analysis. Detailed procedure of this m e thod is descri bed in Standard Methods No. 4500-O B (Eaton et al ., 1995). Various iodom etric m odificati ons have been developed. One procedure that is suitable for use in the field involves using Hach ® high range dissolved oxygen am poules (Hach Com p any, Loveland, CO). Once the Hach ® am poules are filled and capped and the reaction of the reagents with DO in the water sam p le is com p lete, the colo r is stable indef i nitely. Theref ore, the capped
am poules can be transported back to the laborat ory trailer where the DO concentration in the water sam p les can be determ ined with a Hach ® colorim e ter. (Note: the caps often leak gaseous O 2 into the sam p les; so, the sam p les aren’t rea lly stable indefinitely, even though the color is stable as long as additional O 2 can be excluded.)
The m e m b rane electrode m e thod is m o re suitable for regular field m onitoring and in situ DO determ ination. It is also recomm e nded for DO an alysis in highly polluted waters and colored waters. The general procedure for this m e thod is described in Standard Methods No. 4500-OG (Eaton et al ., 1995). Detailed analytical procedures m a y vary depending on the m anufacturers of the DO probes.
3.4.2 pH Biodegradation of oil is affected by th e background pH (Atlas and Bartha, 1992). Oil biodegradation can also be severely inhibited by dram atic reductions in pH when amm onia is provided as the nitrogen source (Wrenn et al ., 1994). The latter is true for closed environm ents such a laboratory flasks where no dilution is possible from continuously changing aqueous conditions such as tides. Nonetheless, m onitoring pH in the f i eld is of particular im portance in evaluating the effect of oil biodegradation. The pH values of pore-water sam p les are norm a llym easured using a portable pH m e ter with a
com b ination electrode. The pH can be m easured inthe field either imm e diately after the sam p les are collected or by putting pH electrodes directly into water in sam p ling wells. Sedim e nt pH can also be m easured in the field by m i xing the so il sam p les with reagent water according to EPA
m e thod 9045c (US EPA, 1992).
3.4.3 Temperature As discussed in Chapter 2, tem p erature affe cts both the properties of spilled oil and the biodegradation processes. All the other m eas urem ents are also tem p erature-dependent. Tem p erature profiles in air, water , and sedim e nt should be m onitored regularly using appropriate

67therm o m e ters. Many instrum e nts for other an alyses, such as DO or pH, have built-in therm o m e ters, and the tem p erature should be r ecorded along with the other m easurem ents. Any tem p erature m easurem ent devices should be calibra ted with a National Institute of Standards and Technology (NIST)-certified therm o m e ter before field use (Eaton et al ., 1995).
3.4.4 Salinity Salinity of the environm ent m ay be an im portant factor in oil biorem ediation, particularly in estuarine environm ents or in m arine shorelines where regional seaward groundwater flow exists (Boufadel, et al ., 1999b; Zobell, 1973). Salinity of pore water can be m easured by either a conductivity m e thod or density m e thods (e.g., Standard Methods No. 2520B or 2520 C; Eaton et al ., 1995). The conductivity m e thod is the
m o st comm only used m e thod for salinity anal ysis due to its high sensitivity and ease of
m easurem ent. Conductivity m e ters can be instal led in the field to m onitor salinity profiles in
m arine shorelines. The density m e thod involves using a precise vibrating flow densitom e ter. W a ter salinity can also be determ ined rapidly in the field using this m e thod.
3.5 Monitoring of Biological Impacts The public has responded favourably to biorem ediati on since its im plicit goal is that of reducing toxic effects by converting organic m o lecules to cell biom ass and other benign m a terials such as carbon dioxide and water (Atlas and Cerniglia, 1995). However, concerns about the net benefit of biorem ediation strategies rem a in. This is attributed to lingering questions regarding the potential production of toxic m e tabolic by-pr oducts, possible toxic com ponents in the form ulation of biorem ediation agents, and the ineffective degradation of the m o st toxic
com ponents of residual oils (Hoff, 1991; Office of Technology Assessm ent, 1991). To date, a single ideal biological m e thod - both sensitive and efficient - for the assessm ent of contam inant im pacts to all sedim e nt biota ha s not been identified. Two separate, yet com p lim entary, approaches have evolved: bi oassessm ent and bioassays. Bioassessm ents are field-based analyses typically characterized by assessing the im pacts of the contam ination and treatm e nt activity on environm ental populations such as benthic comm unities, intertidal flora and fauna, etc. They are characterized as having lim ited experim e ntal controls (Herricks and Schaeffer, 1984). Bioassays are laboratory-based tests that incorporate rigorous experim e ntal protocols and controls. Both toxicity tests a nd bioaccum u lation studies are bioassays (Chapm an, 1989).
3.5.1 Bioassessment Changes in benthic comm unity structure can be used as a m eans of assessing ecosystem response to contam inated sedim e nts in aquatic ecosystem s
. Since m o st contam inants such as crude oil within the aquatic ecosystem eventually bind to sedim e nt particles, em phasis on benthic organism s (bottom dwelling vertebrates and inve rtebrates) as a prim ary m eans of assessing ecosystem response is warranted. Of particular im portance are the m acrobenthic invertebrates (organism s retained on screens of m e sh size >0.2 mm ) because of their basic longevity,

68sedentary lifestyles, proxim i ty to sedim e nts, influence on sedim e ntary processes, and trophic im portance. Microinvertebrates such as rotifers and nem a todes are of particular ecological interest; however, their taxonomy is less well known. Hence, they have not been routinely
m onitored in environm ental assessm ents.
Wh ile there is a vast bioassessm ent database on th e ef f ects of oil spills, the ef f ects of clean up techniques have until recently been seldom addre ssed. Clearly, a database on the effects of clean-up operations would have obvi ous potential for guidance. For exam ple, in a follow up of the Exxon Valdez oil spill clean up, Driskell et al . (1996) noted that total abundance, species richness, species diversity, and abundance of se veral m a jor taxa (polychaetes, bivalves, and gastropods) were significantly lower in hot-wat er-washed beaches than in unoiled beaches. Infauna at oiled sites that were not hot -water washed rebounded quickly following the disturbance. Three years af ter the spill, recovery of inf auna at sites that were cleaned still lagged significantly behind the oiled sites. Principa l com ponent analysis (PCA), a m u ltivariate ordination technique, was used to track site rec overy trends. Negative effects were indicated by reduction in size or biom ass, m o rtality, and reduc ed or failed reproductive success. Conversely, the possibility of positive im pacts was also iden tif ied (e.g., when oil tolerant species bloom during the period of reduced com p etition-predation). Changes in epif auna and inf auna were also used to assess the rates of natural recovery and the im pacts of intertidal clean-up activities on the coast of Saudi Arabia following the 1991 Gulf oil spill (W att et al ., 1993).
Macroinvertebrate bioassessm ent has been lim ited in field trials evaluating the efficacy of biorem ediation strategies due to the am ount of unrestricted surface area required for sam p le collection. The use of bioassessm ents in this context will expand with the developm ent of bioassay protocols based on bioanalytical tec hniques (enyzm a tic m easurem ents, as well as imm unoassay and biosensor techniques) aim e d at the subcellular or m u lticellular level of biological organization. Application of these ki nds of tests should be tailored for both the field and laboratory (Lee et al ., 1998).
Bioassessm ent can readily include potential im pact s on vegetation. Field surveys dem onstrated that the 1991 Gulf W ar oil spill severely dam a ge d intertidal vegetation along the Saudi Arabian Gulf coast (Boer, 1994). Along a 45 km stretch of intertidal m angroves and salt-m arshes, Salicornia europaea was alm o st extinct. Dwarf m angrove ( Avicennia marina ) and salt-m arshes dom inated by Arthrocnemum macrostachyum and Halocnemum strobilaceum were severely dam a ged. Halopeplis perfoliata and Limonium axillare salt-m arshes were relatively unaffected. It was noted that natural re-establishm ent of th e vegetation would be pr otracted unless active
m easures were taken to aid recovery.
A bioassessm ent of vegetative growth was r ecently used to docum ent the efficacy of biorem ediation strategies to enhance the rate of habitat recovery within a tidal f r eshwater m arsh located along the St. Lawrence River, Canada (Lee et al ., 2001). Scirpus pungens, the dom inant plant species at the study site was found to be tolerant to the oil, and its growth was significantly enhanced above that of the unoiled control by the addition of nutrients (Figure 3.5).
The aim of oil spill rem e diation is to restore a site to its pre-spill condition. In this context,
m onitoring the recolonization of im pacted areas should be a prim ary goal in bioassessm ents.

69Colonization and succession describe changes in the num bers and kinds of organism s m a king up the comm unity over tim e. They provide an integrated m e asure of a toxicant’s effect on imm i gration, em igration, com p etition and predation. Colonization is som e what analogous to reproduction in a single species: it reflects the abilityof the comm unity to replicate and organize itself. Fleeger et al . (1996) showed that unweathered Exxon Valdez crude oil delayed, but did not preclude, colonization by m e iof auna (harpactic oids) into azoic sedim e nt of Prince W illiam Sound. To date, em phasis has been placed on the charact erization of im pacts on the m acroinvertebrate
comm unity and vegetation. Nevertheless, se dim e nt-associated contam inants enter the nonbenthic environm ent and comm unity through natural processes including resuspension, desorption, ingestion of benthic organism s, inge stion of sedim e nt, and adsorption to or uptake through m e m b ranes during sedim e nt contact. Due to m obility and sam p ling issues, it is inherently m u ch m o re difficult to work on pelagi c organism s such as fish. Nevertheless, given the holistic nature of toxicant perturbati ons on aquatic ecosystem s and the m u ltifaceted interactions between the water and sedim e nt co m p artm ents, consideration should also be given to the bioassessm ent of fish and other nonbenthic comm unity organism s (e. g., bacteria, phytoplankton, cladocera, and am phibians).

Figure 3.5 Minim um and m a xim um height of the (a) dom inant ( Scirpus pungens) and (b) secondary (Eleocharis palustris) plant species at W eek15. Treatm e nt of the oiled plots included natural attenuation (Nat. Att.); nutrient am endm en t with granular amm onium nitrate and super triple phosphate (NH 4 + Intact); nutrient am endm ent with s odium nitrate instead of amm onium nitrate (NO 3 - ). Unoiled plots were am ended with granular amm onium nitrate and super triple phosphate (NH 4 + Unoiled).

703.5.2 Bioassays Bioassays provide a m o re accurate picture of eco system health at a contam inated site than chem ical analyses because their result is an integr ation of the interaction that occurs between the contam inant and environm ental variables. Bi oassay endpoints are quantitative m e asures of toxicity. They com p lim ent biological surveys, which describe comm unities of organism s in the field, and chem ical analyses, which provide inform ation on the nature of contam inants at a site, the m a gnitude of the rem e diation problem , and potential m e thods of treatm e nt. Resource
m a nagers frequently use bioassays to identify th e m o st toxic areas, thereby helping to prioritize sites for m o re thorough evaluation, including the selection of m e thods for chem ical analyses.
Sedim e nt toxicity tests are generally classifi ed as “acute” or “chronic”. They are usually perform e d on whole sedim e nt (e.g., solid-phase), suspended sedim e nt, sedim e nt liquid phases (pore water, interstitial water), or sedim e nt extract s (elutriates, solvent extracts). In general, assays using whole sedim e nt sam p les are m o re sens itive than assays using elutriate or pore water
sam p les. The Am erican Society of Testing a nd Materials (ASTM, 1991) currently defines an acute toxicity test as a com p arative study in wh ich the organism s that are subject to different treatm e nts are observed for a short period, usuallynot constituting a substantial portion of their life span. A chronic test is defined as a com p ar ative study in which organism s that are subjected to different treatm e nts are observed for a long peri od or a substantial porti on of their life span. Acute tests of ten utilize m o rtality as the only m easure of ef f ect, while chronic tests usually include m easures of growth, m o rphology, reproducti on, behavioral effects, or other sublethal endpoints.
Plant and anim al comm unities are diverse; their m e m b ers dif f e r in their sensitivity to toxicants. A single species bioassay cannot represent the ra nge of sensitivity of all biota within an ecosystem

. To im prove ecological relevance, a test battery approach with species from different trophic levels is required. Accountability for th e influence of natural environm ental factors in sedim e nt bioassays is assisted by the testing of reference and control sam p les. Reference sedim e nt m ay be defined as sedim e nt collected from the vicinity of a study site, possessing sim i lar characteristics to the test sedim e nt, but without anthropogenic contam inants. Sedim e nt characteristics, such as particle size distribu tion and percent organic carbon of the reference sedim e nt should sim u late, as closely as possible, that of the test sedim e nt. In som e cases the reference sedim e nt m i ght also show toxicity due to naturally occurring chem ical, physical, or biological properties. This factor can be a ddressed by determ ining the toxicity of control sedim e nts (natural or artificially prepared sedim e nts known to be nontoxic) and the use of positive controls (a sedim e nt of known toxicity to the test organism under the conditions of the test). Bioassays have been developed and used exte nsively since the 1960s for the screening of chem icals and regulatory com p liance m onitori ng. Sedim e nt bioassays have been used extensively to diagnose the effects of oil spills (Teal, et al ., 1992; Gilfillan et al ., 1995; Neff and Stubblefield, 1995; Randolph et al ., 1998). Their application has now been extended to include the docum entation of effects and success of oil spill counterm easures like biorem ediation (Lee et al ., 1995b; Mearns et al ., 1995). Wh ile any living organism s can be used in theory, toxicity tests with fish and m acroinvertebrates have been standardized by environm ental agencies to assess the

71hazards of industrial wastes to aquatic system s (Blaise et al ., 1988). Rapid advances are now being m a de in the developm ent of cost-effec tive high-perform ance m i cro-scale procedures involving bacterial, protozoan, m i croalgal, and m i croinvertebrate indicators (W ells et al ., 1998). Furtherm ore, the high dem and for sim p le, rapid, and practical toxicological procedures has resulted in the creation of comm e rcial bioanalytical products such as the Microtox  Test (AZUR Environm ental Inc., USA). Major criteria to consider in the selection of species f o r sedim e nt toxicity testing include: (1) their behavior in sedim e nt (habitat, f eeding habits , etc.), (2) their sensitivity to test m a terial, (3) their ecological and/or econom ic relevance, (4 ) their geographical distribution, (5) their taxonom ic relation to indigenous anim als, (6) thei r acceptability for use in toxicity m e asurem ent (standardized test m e thod), (7) their availability , and (8) their tolerance to natural sedim e nt characteristics such as grain size. The response of the test organism s to the toxicant or test sedim e nt is often affected by its life stage. Larval or juvenile lif e stages are generally m o re sensitive than adults.
3.5.2.1 Benthic invertebrates
In term s of benthic invertebrates, am phipods are am ong the m o st sensitive of benthic species. They are am ong the first to disappear from benthic comm unities in sedim e nts im pacted by pollution (Swartz et al ., 1982; Mearns and W o rd, 1982). They have been used successfully to characterize shoreline im pacts following oil spill incidents (Teal et al . 1992; Gilfillan et al ., 1995; W o lfe et al ., 1996).
Gilf illan et al . (1995) collected m u ssels from several lo cations for tissue hydrocarbon analysis to estim ate bioavailable hydrocarbon concentrations in epifaunal species. In a newer approach, studies of sedim e nt contam ination and verificat ion of laboratory bioassays involved controlled in situ exposures (caged anim als) to expand the level of ecological relevance. In this case, oysters were used during a shoreline biorem ediation expe rim e nt in Delaware Bay to docum ent the loss of oil from the study area and to determ ine how the overall oiling m ay have im pacted offshore resources (Mearns et al ., 1997).
3.5.2.2 Microtox
Sim p le, sensitive, rapid, cost-effective, reproducible , and practical m e thods are needed for the screening of toxic im pacts during oil spill response operations. The Microtox  Test, a comm e rcial bioassay accepted by regulatory agencies, is based on the m easurem ent of changes in light em ission by a nonpathogenic, biolum inescent m a rine bacterium ( Vibrio fisheri ) upon exposure to test sam p les. The test has been us ed extensively worldwide over the last 18 years for toxicity screening of chem icals, effluents, wa ter and sedim e nt, and for contam ination surveys and environm ental risk assessm ent. Variations of this test have been applied to tim e-series
m onitoring of sedim e nt and water toxicity. Ho and Quinn (1993) identified strong rank correlations between the Microtox response and polyc yclic arom atic fractions of organic extracts of sedim e nts. Its application for m onitoring th e efficacy of oil spill rem e diation m e thods has

72been proven (Lee et al. , 1995b; 1997b). Mueller et al . (1999) quantified the effectiveness of intrinsic recovery within an oiled wetland by m onitoring the rate of acute toxicity reduction using the Microtox 100% Test on a water extracta ble phase. The observed decrease in toxicity followed a pattern sim ilar to the decrease seenin petroleum concentrations by GC/MS total target analyte m easurem ents.
3.5.2.3 Fish
Due to their econom ic, recreational, and aesthetic valu e, fish have been historically selected as a prim ary bioassay organism

. Dif f i culties in using f i sh as biom onitors of sedim e nt contam ination arise from their preference for particular sedim e nts or habitats and their residence tim e in or over contam inated areas. Furtherm ore, their absence in a water body m a y m o re directly reflect water quality.
Biochem i cal and physiological alterations, if severe enough or protracted, can lead to structural alterations in organelles, cells, tissue, and or gans. Detection of specific alterations using anatom ical and cytological endpoints m ay indicate both prior and current exposure to chem ical contam inants, so histopathology has been instrum e ntal in assessing the toxicological im pact of contam inated sedim e nts. The docum entation of neoplasm s in fish and other aquatic organism s was perhaps the first use of histopa thological indices in ecotoxicology.
Biom arkers (as distinguished from the oil biom arkers such as hopane discussed earlier) are used by resource m a nagers as a m e ans to identify a toxicological response from fish populations. Biom arkers can be defined as biochem i cal, physio logical, or pathological responses m easured in individual organism s on exposure to environm en tal contam inants, and which also provide inform ation concerning sublethal effects arising from such exposures. The fam ily of enzym e s referred to as cytochrom e P-450s (or P450) act on the functional groups of lipophilic substrates in a process referred to as m i xed function oxida se (MFO) or m onooxygenase reactions (Ortiz de Montellano, 1986). MFO reactions are induced by polycyclic arom atic hydrocarbons (PAHs) and a variety of halogenated hydr ocarbons (notably certain chlorinated biphenyls, dibenzofurans , and dibenzodioxins). The enzym e system is sensitive to these contam inants at levels encountered in the environm ent. In fish, th e m o st widely em ployed and readily perform e d techniques are m easurem ents of enzym e activ ities, particularly aryl hydrocarbon hydroxylase (AAH) and ethoxyresorufin O-deethylase (EROD), th at are highly associated with the P450 1A proteins.
Hodson et al . (2001) has m onitored changes in the bioa vailability and toxicity of oil-derived PAH to early life stages of fish in a field trial to evaluate the ef f ectiveness of wetland biorem ediation and phytorem ediation strategies

. For over 1.5 years, sedim e nts from experim e ntal plots were tested by bioassays of MFO (CYP1A) enzym e activity in livers of trout as an index of PAH exposure. Oil alone, oil m i xed with sedim e nts in the lab, and oiled sedim e nts from the experim e ntal plots all cause d CYP1A induction relative to unoiled controls, indicating the presence and bioavailability of PAH. Induction did not vary m a rkedly am ong treatm e nts, but declined slowly with tim e. Conc om itant chem ical analysis suggested that PAHs were depleted prim arily by weathering or sedi m e nt dispersion rather than by biorem ediation

73treatm e nts. Sedim e nts were also chronically t oxic to developing stages of trout and m e daka ( Oryzias latipes ), causing increased rates of def o rm ities and m o rtality.
3.5.3 Application of bioassays to assess bioremediation in marine environments Bioassays were used to docum ent the effectiveness of shoreline biorem ediation in accelerating toxicity reduction of a sandy shoreline at Fowler Beach, Delaware, USA, that had been oiled with weathered Nigerian Bonny light crude (Mearns et al ., 1995). The bioassay suite included two solid phase (Am phipod Survival, Microtox So lid-Phase) and three pore water tests (Grass Shrim p Em bryo Survival/Growth, Microtox, Sea Urch in Fertilization). Treatm e nt with nutrients (sodium nitrate and sodium tripolyphosphate) or nutrients and oil degrading bacteria (isolated from the study site) did not accelerate toxicity reduction. However, results of the high-frequency test based on the hatching success of grass shri m p em bryo suggested there m ay have been a substantial delay in pore water toxicity reduction due to the addition of the nutrients them selves during the first few weeks. The Sea Urchin Fert ilization Test was least sensitive. The m o st sensitive tests were the 10-day am phipod and grass shrim p em bryo bioassays.
Biorem ediation by nutrient enrichm e nt was inves tigated as a m e thod of treating a m i xture of Forties Crude Oil and Heavy Crude Oil stranded on Bullwell Bay, Milford Haven, UK, after the grounding of the Sea Em press in 1996 (Swannell et al ., 1999b). Experim e ntal results showed that the oil was significantly m o re biodegraded afte r two m onths as a result of application of the fertilizer. Based on the results of a bioassay th at involved m onitoring the developm ent of oyster
em bryos, and the results of the Microtox Organic Solvent Basic Test, there was no evidence of detrim ental effects associated with the biorem ediation treatm e nts. To date, detrim ental effects from nutrient enrich m e nt have not been observed following actual field operations (Prince, 1993; Mearns et al ., 1997) although the possibility of a future incident still exists. As an exam ple, oxygen depleti on and production of amm onia from excessive applications of a fish-bone m eal fertilizer duri ng one field experim e nt caused detrim ental effects that included toxicity and the suppr ession of oil degradation rates (Lee et al ., 1995b). Furtherm ore, in a subsequent biorem ediation f i eld trial it was reported that a comm e rcial biorem ediation product suppressed the rates of toxi city reduction as it increased the retention of residual oil within the sedim e nts (Lee et al ., 1997b). For safety assurance, future operational guidelines m u st include ecotoxicological m onitoring protocols.

74Chapter 4 TYPES OF AMENDMENTS AND CONSIDERATIONS IN THEIR APPLICATION The success of oil spill biorem ediation depends on our ability to optim ize various physical, chem ical and biological conditions in the contam inated environm ent. Existing am endm ents for enhancing oil biodegradation in m arine shorelines and freshwater wetlands include addition of nutrients, addition of m i crobial cultures or enzym e s, phytorem ediation, and oxygen enhancem ent.
4.1 Nutrient Amendment As reviewed in Chapter 2, nutrient addition ha s been proven to be an effective strategy to enhance oil biodegradation in various m arine shorelines. Theoretically, approxim a tely 150 m g of nitrogen and 30 m g phosphorus are consum ed in the conversion of 1 g of hydrocarbon to cell
m a terial (Rosenberg and Ron, 1996). Therefore, a comm only used strategy has been to add nutrients at concentrations that approaches a st oichiom e tric ratio of C:N:P of 100:5:1. Recently, the potential application of resource-ratio theo ry in hydrocarbon biodegradation was discussed (Head and Swannell, 1999; Sm ith et al ., 1998). This theory suggests that m anipulating the N:P ratio m ay result in the enrichm e nt of different m i crobial populations, and the optim al N:P ratio can be different for degradation of different com pounds (such as hydrocarbons m i xed in with other biogenic com pounds in soil). However, the practical use of these ratio-based theories
rem a ins a challenge. Particularly, in m arine shorelines, m a intaining a certain nutrient ratio is im possible because of the dynam i c washout of nutrients resulting from the action of tides and waves. A m o re practical approach is to m a intainthe concentrations of the lim iting nutrient or nutrients within the pore wate r at an optim al range (Bragg et al ., 1994; Venosa et al ., 1996). Comm only used nutrients include water-soluble nutrients, solid slow-release nutrients, and oleophilic f e rtilizers. Each type of nutrien t has its advantages and lim itations. General characteristics of these nutrients and im portant factors affecting their persistence in the field, such as waves and tides, and physical intrusion e ffects, will be discussed in this section and
summ arized in Table 4.1. More practical issues su ch as nutrient application strategies will be discussed in Chapter 5.
4.1.1 Water-soluble nutrients
Comm only used water-soluble nutrient products include m i neral nutrient salts (e.g. KNO 3 , NaNO 3 , NH 3 NO 3 , K2 HPO 4 , MgNH 4 PO 4 ), and m any comm e rcial inorganic f e rtilizers (e.g. the 23:2 N:P garden f e rtilizer used in Exxon Valdez case). They are usually applied in the field through the spraying of nutrient solutions or spread ing of dry granules. This approach has been effective in enhancing oil biodegradation in m any field trials (Swannell et al ., 1996; Venosa et al ., 1996). Com p ared to other types of nutrients , water-soluble nutrients are m o re readily available and easier to m anipulate to m a intain ta rget nutrient concentrations in interstitial pore water. Another advantage of this type of nutri ent over organic f e rtilizers is that the use of inorganic nutrients elim inates the possible co m p etition of carbon sources . The field study by Lee et al. (1995a) indicated that although organic fertilizer s had a greater effect on total heterotrophic
m i crobial growth and activity, the inorganic nutrien ts were m u ch m o re ef f e ctive in stim ulating crude oil degradation.

75However, water-soluble nutrients also have several potential disadvantages. First, they are m o re likely to be washed away by the actions of tides and waves because of their water-solubility. The field study in Maine dem onstrated that water-solubl e nutrients can be washed out within a single tidal cycle in high-energy beaches (Wrenn, 2000, s ee section 2.6.2). Second, inorganic nutrients,
amm onia in particular, should be added carefully to avoid reaching toxic levels. Existing field trials, however, have not observed acute toxicity to sensitive species resulting from the addition of excess water-soluble nutrients (Mearns et al ., 1997; Prince et al ., 1994). Third, water-soluble nutrients m ay have to be added m o re frequently than slow release nutrients or organic nutrients, resulting in m o re labor-intensive, costly, and physical intrusive applications.
4.1.2 Granular nutrients (slow

- release) Many attem p ts have been m a de to design nutri ent delivery system s that overcom e the washout problem s characteristic of intertidal environm ents (Prince, 1993). Use of slow release fertilizers is one of the approaches used to provide con tinuous sources of nutrients to oil contam inated areas. Slow release f e rtilizers are norm a lly in so lid f o rm s that consist of inorganic nutrients coated with hydrophobic m a terials lik e paraffin or vegetable oils. This approach m a y also cost less than adding water-soluble nutrients due to less frequent applications. Slow release fertilizers have shown som e prom ises from oil biorem ed iation studies and applications. For exam ple, Olivieri et al .(1976) found that the biodegradation of a crude oil was considerably enhanced by addition of a paraffin coated MgNH 4 PO 4 . Another slow-release f e rtilizer, Custom blen (vegetable oil coated calcium phosphate, amm onium phosphate , and amm onium nitrate), perform e d well on
som e of the shorelines of Prince W illiam Sound, particularly in com b ination with an oleophilic fertilizer (Atlas, 1995a; Pritchard et al ., 1992; Swannell et al ., 1996). Lee et al . (1993) also showed that oil biodegradation rates increased withthe use of a slow release fertilizer (sulfurcoated urea) com p ared to water-soluble fertilizers. However, the m a jor challenge for this technology is control of the release rates so that optim al nutrient concentrations can be m a intained in th e pore water over long tim e periods. For exam ple, if the nutrients are released too quickly, they w ill be subject to rapid washout and will not act as a long-term source. On the other hand, if they are released too slowly, the concentration will never build up to a level that is sufficient to support rapid biodegradation rates, and the resulting stim ulation will be less ef f ective than it could be. The f i eld trials on of the shorelines of Prince W illiam Sound showed that on certain beaches, Custom blen granules were apparently washed away before any significant enhancem en t of biorem ediation was recorded ( Swannell et al ., 1996). Several recent studies have shown that a slow release nutrient (Max Bac, a product sim i lar to Custom blen) failed to dem onstrate enhancem ent of oil degradation because the nutrient release rate was too slow to affect oil biodegradation (Croft et al ., 1995; Sveum and Ram s tad, 1995).
4.1.3 Oleophilic nutrients Another approach to overcom e the problem of wa ter-soluble nutrients being rapidly washed out was to utilize oleophilic organic nutrients (Atla s and Bartha, 1973; Ladousse and Tram ier, 1991). The rationale f o r this strategy is that oil bi odegradation m a inly occurred at the oil-water

76interface; since oleophilic fertilizers are able to adhere to oil and provide nutrients at the oilwater interface, enhanced biodegradation should result without the need to increase nutrient concentrations in the bulk pore water. A well- known oleophilic fertilizer is Inipol EAP 22, a
m i croem u lsion containing urea as a nitrogen source, lauryl phos phate (the phosphorus source), 2- butoxy-1-ethanol as a surfactant, and oleic acid to give the m a terial its hydrophobicity. This fertilizer has been subjected to extensive studies under various shoreline conditions and was successfully used in oil biorem ediation on of the shorelines of Prince W illiam Sound. Other oleophilic organic fertilizers include polym er ized urea and form aldehyde, and som e organic fertilizers derived from natural products such as fishm eal (Lee et al., 1995a; Rosenberg et al ., 1992; Sveum and Ram s tad, 1995).

Table 4.1 Major nutrient types used in oil biorem ediation Type of nutrients Advantages
W a ter soluble Readily available Rapidly washed out Easy to m a nipulate by wave and tide for target nutrient Labor-intensive, and concentrations physical intrusive No com p licated effect applications of organic m a tter Potential toxic ef f ect
Slow release Provide continuous Maintaining optim al sources of nutrients nutrient release rates and m ay be m o re cost could be a challenge effective than other types of nutrients Oleophilic Able to adhere to oil Expensive Disadvantages

Effectiveness is variable Containing organic carbon, which m ay com p ete with oil degradation and result in undesirable anoxic conditions Applications in the field or field trials Alaska (Pritchard et al. , 1992) Delaware (Venosa et al. , 1996)

Alaska (Pritchard et al. , 1992) Nova Scotia (Lee et al ., 1993)

Alaska (Pritchard et al. , 1992) Nova Scotia (Lee et al ., 1987, 1989,1995a &b) and provide nutrients at the oil-water interface

The ef f ectiveness of oleophilic f e rtilizers also de pends on the characteristics of the contam inated environm ent such as action of wave and tide, and sedim e nt types. Based on several earlier studies, Sveum et al. (1994) indicated that oleophilic fertilizers proved to be m o re effective than water-soluble f e rtilizers when the spilled oil reside d in the intertidal zone. But they have no advantages in enhancing oil biodegradation in the supralittoral zone where water transport is lim ited. Inipol EAP 22 was found to be m o re effectiv e in coarse sedim e nts than in fine sedim e nts due to the dif f i culty in penetration f o r the ole ophilic f e rtilizer in f i ne sedim e nts (Sveum and

77Ladousse,1989). Variable results have also been produced regarding the persistence of oleophilic fertilizers. Som e studies showed that Inipol EAP 22 can persist in a sandy beach for a long tim e under sim u lated tide and wave actions (Santas and Santas, 2000; Swannell et al . 1995). Others found that Inipol EAP22 was rapidly washed out before becom i ng available to hydrocarbondegrading bacteria (Lee and Levy, 1987; Safferm an, 1991). Another disadvantage with oleophilic fertilizers is that they contai n organic carbon which m ay be biodegraded by
m i croorganism s in preference to petroleum hydrocarbons (Lee et al ., 1995a; Swannell et al ., 1996), and m ay also result in unde sirable anoxic conditions (Lee et al ., 1995b; Sveum and Ram s tad, 1995).
In summ ary, the ef f ectiveness of these va rious types of nutrients will depend on the characteristics of the contam inated environm ent. Slow-release fertilizers m ay be ideal nutrient sources if the nutrient release rates can be we ll controlled. W a ter-soluble f e rtilizers are likely
m o re cost-effective in low-energy and fine-grained shorelines where water transport is lim ited. And oleophilic f e rtilizers m ay be m o re suitabl e f o r use in high-energy and coarse-grained beaches. However, successful application of biorem ediation products will always require appropriate testing and evaluation based on the specific conditions of each contam inated site.
4.2 Microbial Amendments Addition of oil-degrading m i croorganism s or bioaugm entation has been proposed as a biorem ediation strategy. The rationale for this approach includes the contention that indigenous
m i crobial populations m ay not be capable of degr ading the wide range of substrates that are present in com p lex m i xtures such as petrol eum and that seeding m ay reduce the lag period before biorem ediation begins (Forsyth et al ., 1995; Leahy and Colwell, 1990). For this approach to be successful in the field, the seed m i croorganism s m u st be able to degrade m o st petroleum
com ponents, m a intain genetic stability and viability during storage, survive in foreign and hostile environm ents, effectively com p ete with indige nous m i croorganism s, and m ove through the pores of the sedim e nt to the contam inants (Atlas, 1977; Goldstein et al ., 1985).
There are m any vendors of biorem ediation products , who claim their product (m ost of them are m i crobial agents) aids the oil biodegradation process. The U.S. EPA has com p iled a list of biorem ediation agents (USEPA, 2000) as part of the National Oil and Hazardous Substances Pollution Contingency Plan (NCP) Product Schedule, which is required by the Clean W a ter Act, the Oil Pollution Act of 1990, and the Nationa l Contingency Plan. A current list of biorem ediation agents in NCP schedule is shown inTable 4.2. A product can be listed only when its safety and effectiveness have been dem ons trated under the conditions of a test protocol developed by EPA (NETAC, 1993). However, li sting does not m e an that the product is
recomm e nded or certified for use on an oil spill (U SEPA, 2000). The efficacy test protocol uses laboratory shake flasks to com p are the degradati on of artificially-weathered crude oil in natural seawater with and without a biorem ediation pr oduct. Sim i lar test protocols for freshwater conditions were recently proposed (Haines et al ., 1999).

78Table 4.2 Biorem ediation agents in NCP pr oduct schedule (Adapted from USEPA, 2000) Type Nam e or Tradem ark Manufacture Biological additives (Microbial Culture or Enzym e additives) BET BIOPETRO
BIOGEE HC BR (form e rly ENVIROZYME BR)
ENZYT (LIQUID/CRYSTA)
MICRO-BLAZE
OPPENHEIMER FORMULA
PRISTINE SEA II PRP ( Petroleum Rem e diation Product)
STEP ONE

SYSTEM E.T. 20

W M I-2000 BioEnviro Tech, Tom b all, TX
RMC Biorem ediation, Shreveport, LA Enviro-Zym e, Inc., Stormv ille, NY

Acorn Biotechnical Corporation Houston, TX
Verde Environm ental, Inc., Houston, TX Oppenheim er Biotechnology, Inc. Austin, TX Marine System s, Baton Rouge, LA
Petrol Rem , Inc., Pittsburgh, PA

B & S Research, Inc. Em barrass, MN Quantum Environm ental Technologies, Inc. (QET),
W a ste Microbes, Inc., Houston, TX
Nutrient additives INIPOL EAP 22 (oleophilic)
LAND AND SEA RESTORATION
OIL SPILL EATER II

VB591 TM WA TER, VB997 TM SOIL, AND BINUTRIX (partially encapsulated Societe, CECA S.A. France Land and Sea Restoration LLC, San Antonio, TX Oil Spill Eater International, Corporation Dallas, TX BioNutraTech, Inc., Houston, TX La Jolla, CA

& oleophilic)

79As reviewed in Chapter 2, however, even though the addition of m i croorganism s m ay be able to enhance oil biodegradation in the laboratory, its effectiveness has not been convincingly dem onstrated in the f i eld. Actually, m o st f i el d studies indicated that bioaugm entation is not effective in enhancing oil biodegradation in m arine shorelines, and nutrient addition or biostim ulation alone had a greater effect on o il biodegradation than the m i crobial seeding (Jobson et al ., 1974; Lee and Levy, 1987; Lee et al ., 1997b, Venosa et al ., 1996). The failure of bioaugm entation in the field m ay be attributed to the fact that the carrying capacity of m o st environm ents is likely determ ined by factors th at are not affected by an exogenous source of
m i croorganism s (such as predation by protozoans, the oil surface area, or scouring of attached biom ass by wave activity), and that added bacteria seem to com p ete poorly with the indigenous population (Tagger et al ., 1983; Lee and Levy, 1989; Venosa et al ., 1992). Therefore, it is unlikely that externally added m i croorganism s wi ll persist in a contam inated beach even when they are added in high num bers. In short, t hose criteria m e ntioned above for a successful colonization are very difficult to be m e t in the field. Fortunately, oil-degrading m i croorganism s are ubi quitous in the environm ent, and they can increase by m any orders of m a gnitude after bei ng exposed to crude oil (Atlas, 1981; Lee and Levy, 1987, Pritchard and Costa, 1991). Therefore, in m o st environm ents, there is usually no need to add hydrocarbon degraders. In certain circ um stances that have not been well defined, when the indigenous bacteria are incapable of degrading one or m o re im portant contam inants, addition of m i crobial inocula m ay be consider ed. Genetically engineered organism s are not likely to be used in the near or even distant future.
4.3 Plant Amendments (phytoremediation) Phytorem ediation has been defined as the us e of green plants and their associated
m i croorganism s to degrade, contain, or render harm less environm ental contam inants (Cunningham et al. , 1996). This technique is em erging as a potentially cost-effective option for clean-up of soils contam inated wi th petroleum hydrocarbons (Frick et al. , 1999). As
summ arized by Macek et al . (2000), the m a in advantages of phytorem ediation include less disruption to the environm ent, potential to tr eat a diverse range of contam inants, and high probability of public acceptance. Major concerns regarding this technology include dissolution and m i gration of contam inants, lim itation by the toxicity of the contam inated environm ents, and it being a relatively slow process. Phytorem ediation has been studied in a freshwater environm ent in Quebec, Canada (Garcia-Blanco et al., 2000; Venosa et al., 2002 (subm itted)). These researchers found that addition of nutrients did not result in enhancem ent of biodegradation of crude oil contam inating the plots, whether or not plants were lef tintact or
rem oved. It appeared that in a wetland envir onm ent, oxygen becam e lim iting at depths within a few mm from the surface.
4.3.1 Mechanisms of phytoremediation Phytorem ediation of petroleum hydrocarbons gene rally involves three m a jor m echanism s : (1) degradation, (2) containm ent and (3) the transfer of contam inants f r om soil to the atm o sphere (Cunningham et al. , 1996; Frick et al. , 1999).

80Degradation can be accom p lished by both plants and their associated m i croorganism s. One of the
m o st im portant processes involved in the degrad ation is the interaction between plants and
m i croorganism s in the rhizosphere (root zone). Plants can stim ulate the growth and m e tabolism of soil m i croorganism s by providing root exudate s of carbon, enzym e s, nutrients, and oxygen, which can result in m o re than 100-fold increase in m i crobial counts (Macek et al ., 2000). This process is also m u tual beneficial. The m i crobes can reduce the phytotoxicity of contam inants so that plants can grow in adverse soil conditions. Co m e tabolism m ay also play an im portant role in phytodegradation. Ferro et al. (1997) suggested that plant exuda tes m i ght have served as com e tabolites in enhancing the biodegradation of pyrene in the rhizosphere.
Other m a jor m echanism s of phytorem ediation in clude containm ent of petroleum hydrocarbons and their transfer from the soil to the atm o sphere. Containm ent involves the accum u lation of contam inants within the plants, adsorption of contam inants onto roots, and binding of contam inants in the rhizosphere through enzym a tic activities (Cunningham et al. , 1996; Frick et al. , 1999). Plants can also transport volatile pe troleum hydrocarbons to the atm o sphere through leaves and stem s. However, these effects are le ss im portant than the degradation m e chanism in phytorem ediation of petroleum hydrocarbons (Ferro et al. 1997).
4.3.2 Considerations in application of oil phytoremediation
Wh ile phytorem ediation had been used successful ly within the terrestrial environm ent to decontam inate soils (Banks and Schwab, 1993; Schnoor et al. , 1995), the technique has not been
em ployed as an operational oil spill counterm e asur e. Until recently, only lim ited research had been carried out on the effectiveness of phytor em ediation in freshwater wetlands (Lin and Mendelssohn, 1998). Most of the studies were gr eenhouse experim e nts rather than field studies. Like bioaugm entation, studies on phytorem ediation of petroleum hydrocarbons have produced
m i xed results. The effectiveness of phytorem ediation is site-specific, which can be affected by factors as oil properties, types of plants, and environm ental conditions.
Oil concentrations Plants can tolerate oils with certain concentrat ion ranges. Wh en oil concentrations are too high, toxic ef f ects will lead to growth inhibition or de ath of plants. Wh en oil concentrations are too low, phytorem ediation will not be effective either due to poor bioavailability. Longpre et al. (1999) investigated the im pact of oil con centrations on a freshwater wetland plant ( Scirpus pungens) along the shore of the St. Lawrence River

. The results showed that the plant growth was stim ulated in the presence of crude oil at a concentration less than 4.56g/Kg sedim e nt when
com p ared to the growth of the control plan ts. At higher oil concentrations, up to 27.4g/kg sedim e nt, growth inhibition or no growth increas e was observed. Wh en the oil concentration was above 36.4 g/kg sedim e nt, plant growth was signi ficantly reduced. The study concluded that the plants were likely to survive and grow in sedim e nts contam inated with crude oil in a range of concentrations com p arable to oil spill incidents. Plant species Another im portant factor in considering phytorem ed iation is establishm ent of appropriate plants. Lin and Mendelssohn (1996) studied the effect of oil spills on four freshwater m arsh plant species. Two of them ( C. ordoratus , and A. teres ) f a iled to survive in any of the oiled sods andE.

81quadrangulata could only persist at oil levels up to 8 L/m 2 . In contrast, the growth of Sagittaria lancifolia was enhanced in response of oil addition up to 24 L/m 2

. Generally, legum e and grass species have been the choices for their pot ential use in phytorem ediation of petroleum hydrocarbons (Frick et al. , 1999). Legum es are nitrogen-fi xing plants, which m ay have advantages in com p eting with non-legum e species in oil-contam inated sedim e nts. Native plants should norm a lly be selected since they have better chance to survive or out-compete nonindigenous inocula (Cunningham et al., 1996). Environmental factors Sim ilar to the environm ental factors affecting m i crobial biodegradation discussed in chapter 2,
m a jor environm ental factors affecting phytorem ediation of petroleum hydrocarbons include soil types, nutrients, oxygen and tem p erature. Deta iled description can be seen in Cunningham et al. , 1996 and Frick et al. , 1999.
4.3.3 Applications in marine shoreline and freshw ater w e tlands Current applications of phytorem ediation in m ari ne shorelines and freshwater wetlands have been lim ited to accelerate recovery and restora tion of oiled wetland. For exam ple, m angroves were successfully replanted to restore oil-k illed m angrove forest in Panam aafter the 1986 Refineria Panama oil spill (Teas et al ., 1989). Only a few field studi es have been carried out on the effectiveness of phytorem ediation in enhanci ng oil degradation in m arine shorelines and freshwater wetlands. Lin and Mendelssohn (1998) investigated the eff ects of biostim ulation and phytorem ediation in enhancing habitat restoration and oil degrada tion in a coastal wetland environm ent (greenhouse study). They found that application of fertilizer in conjunction with the presence of transplants led to m u ch higher oil degradation rates than phyt orem ediation alone. The results were attributed to a higher m i crobial num ber and activity induced by the f e rtilizer. However, it was still not clear whether this effect was due to biostim ulation of soil m i croorganism s or due to phytorem ediation via fertilizer-increased plant biom ass. Thes e confounding effects perhaps could have been distinguished by adding one m o re treatm e nt (bio stim ulation with absence of transplants). In 1999 and 2000, a m a jor research study was conduc ted on the shoreline of the St. Lawrence River (Garcia-Blanco et al. , 2000; Venosa et al. , 2002 (subm itted)). The experim e ntal design was sim ilar to the one used on the m arine shoreline in Delaware Bay (Venosa et al. , 1996). There were 5 treatm e nts: a no oil control and four o iled treatm e nts. The oiled treatm ents included a natural attenuation control plot with no am endm ents, a plot receiving amm onium nitrate and orthophosphate nutrients but with the wetland plants continually cut back to ground surface to suppress photosynthetic activity and growth, a plot receiving the sam e nutrients as Treatm e nt B but with the plants left intact, and a plot si m i lar to Treatm e nt C but with only nitrate (no
amm onium ) serving as the nitrogen source. The no-oi l control also received the sam e nutrients as the oiled treatm e nts receiving nutrients. Findings are summ arized as follows: (1) alkane degraders increased only m arginally in all treatm e nts while the PAH degraders were stim ulated to increase by 3.5 orders of m a gnitude in respons e to exposure to crude oil; (2) nitrogen in the form of amm onium was partly adsorbed to negati vely charged soil particles, partly taken up by the root system of the wetland plants, and partlyleached into the pore water. Nitrogen in the

82form of nitrate leached into th e pore water, and som e was taken up in the root system (Lee et al ., 2002 (subm itted); (3) the prim ary m echanism of oil m a ss loss from all the plots, regardless of treatm e nt, was physical rather than biological; (4) with respect to biodegradation of total alkanes and PAHs during the first 21 weeks of the inve stigation as m easured by GC/MS analysis, only about 35% biodegradation occurred in all treatm e nts on average, and no significant differences
am ong any of the treatm e nts were observed (p > 0.05) ; (5) a substantial increase in plant biom ass was observed due to fertilizer addition; (6) bette r biodegradation occurred in surface sam p les in plots where the plants had been rem oved than in any of the core sam p les because of the oxic nature of the surface and the lack of com p etiti on for nutrients by the plant species. Enhanced oxygen transfer to the rhizosphere by the plants th rough their roots did not appear to take place, at least at the level needed by hydrocarbon degraders to m e tabolize the oil rapidly.
The m a jor reason for the lack of biodegradation beyond only about 35% was ascribed to the fact that the oil had been raked into the top 2-3 cm of sedim e nt to m a ke sure that penetration had occurred. Wh en such oil penetration occurs, little oxygen is available to allow significant biodegradation to take place throughout the oiled zone. If oil contam inates only the surface where m o re aerobic conditions exist, and if it doe s not penetrate deeply into the subsurface, better biodegradation should take place, at leas t theoretically. The m a jor conclusion reached from this study was that biorem ediation of an oil-contam inated freshwater wetland where significant penetration of oil has taken place into the sediment has lim ited potential for enhanced cleanup of the contam ination.
In summ ary, the effectiveness of phytorem ediation in enhancing oil degradation in freshwater wetlands is highly site-specific and does not prom ise to be an effective oil cleanup technique. However, it does show prom ise in accelerati ng the recovery and restoration of wetland environm ents contam inated with oil and oil products.
4.4. Oxygen Amendment Oxygen usually is not a lim iting factor on m any sandy beaches. However, oxygen lim itation m ay occur in wetlands and fine-grained shorelines as indicated by som e field studies (Garcia-Blanco et al. , 2001b, Lee and Levy, 1991, Purandare, 1999). Under such circum stances, oxygen am endm ent m ay be considered as a biorem ediation strategy. Although oxygen supply has been widely used for biorem ediation of oil contam inated soils and groundwater, such as at m any subsurface fuel contam inated sites, this strategy has not been applied to enhance oil degradation in m a rine shoreline and freshwater wetlands

. This is because oxygen am endm ent usually involves expensive and environm entally intrusive operations. For biorem ediation of a large-scale oil spill, use of this approach is probably not practical even when oxygen is a lim iting factor. However, under certain circum stance that involves high oil contam ination in sm aller scale and in less sensitive habitats, oxygen am endm ent can still be considered as an alternative for oil biorem ediation. Comm only used oxygen supply tec hniques include tilling, forced aeration, and chem ical m e thods (Atlas, 1991; Brown and Cros bie, 1994; Riser-Roberts, 1998). These m e thods
are summ arized below. They are m o stly based on studies and practices in soil environm ents. Special attention will be given to their potentia ls of application in m arine shorelines and freshwater wetlands.

834.4.1 Tilling Tilling has been a conventional physical m e thod to accelerate natural oil rem oval by exposing oiled sedim e nts to a higher level of physical abrasion and biochem i cal degradation (Owens, 1998). This technique is also an effective m e ans of aeration for surface layer of sedim e nts. It has been successfully used to accelerate biodegr adation in landfarm i ng (Atlas, 1991; Jerger et al ., 1994). Traditional tilling m achines, such as disk harro ws and rototillers, can aerate surface soils to a depth of 6 to 24 inches. Sedim e nts deeper than about 2 ft (60 cm ) can be aerated by using construction equipm ent, such as a backhoe (Riser-Roberts, 1998). Currently, tilling has been recomm e nded as a phys ical m e thod to accelerate natural weathering processes of oil in sandy or coarse-sedim ent beaches (Owens, 1998). The m a in purpose of this practice is to increase physical abrasion of oils rather than to enhance aeration since oxygen is usually not a lim iting factor in these environm ents. However, based on existing experiences in landfarm i ng, this technique m ay have som e poten tial in enhancing oil biodegradation in som e fine-sedim ent beaches where oxygen is lim ited. Tilling is also considered a low-cost technology
am ong the available aeration m e thods (Jerger et al ., 1994). Major concerns regarding this technique incl ude disturbance of both the natural shape of shorelines and local habitats and the potential of releasing of oil and oiled sedim e nt into adjacent locations. The experience from the St. Lawrence Ri ver field trial also suggests that the tilling of surface soil m ay cause oil penetration deep into the shoreline sedim e nts and m ay reduce the overall oil biodegradation rates if the oil penetr ates into anaerobic sedim e nts (Garcia-Blanco et al ., 2001; Venosa et al. , 2002 (subm itted); see Section 5.5.2).
4.4.2 Forced aeration Forced aeration techniques, including injecti on of aerated water, air and pure oxygen, are expensive m e thods and comm only used for enhancing biorem ediation in subsurface sedim e nts and groundwater contam inated with petroleum hydrocarbons (Brown and Crosbie, 1994; RiserRoberts, 1998). Oil contam ination of coastal and wetland environm ents, however, usually occurs near the surface, especially when the contam ination is the result of an offshore spill. Furtherm ore, crude oil rarely penetrates coastal sedim e nts to depths of greater than one foot (Gundlach, 1987). Therefore, these techniques of subsurface aeration are probably not appropriate for use in biorem ediation of oil spill in m arine shorelines and freshwater wetlands.
4.4.3 Chemical methods
Chem ical m e thods involve addition of altern ative oxygen sources such as hydrogen peroxide (H 2 O 2 ), or alternative electron acceptors such as nitrate. Hydrogen peroxide can provide oxygen at a rate up to two orders of m a gnitude faster than the forced aeration m e thods (Brown and Norris, 1994). It also requires less equipm ent a nd capital cost. However, problem s including too rapid decom position, gas blockage, and ineffici ent use were encountered in som e sites when using H 2 O 2 (Brown and Norris, 1994). The chem ical also can be toxic to m i croorganism s at high concentrations (Riser-Roberts, 1998).

84Nitrate has received m o st attention as an alternative electron acceptor because it is relatively inexpensive, very soluble in water, and does not decom pose. Since nitrate is also comm only used nutrient source for oil biostim ulation, addition of nitrate m ay be a prom ising option for oil biorem ediation under oxygen lim iting conditions. A pot ential disadvantage of this m e thod is that nitrate m ay be effective for degr adation of fewer classes of com pounds than oxygen. It has been reported that nitrate-utilizing bacteria can de grade m any arom atics but do not degrade aliphatic
com pounds and benzene under denitrifying conditions (Brown et al. , 1993). However, others recently found that degradation of alkanes coul d take place under denitrifying conditions (Hess et al ., 1996). Sulfate is another potentially useful el ectron acceptor especially in certain m a rine environm ents, such as salt m arshes, where sulf ate reduction is one of m o st im portant natural processes (Mitsch and Gosselink, 2000). Som e laboratory studies have shown that PAHs and alkanes can be degraded under sulfate-reducing c onditions at sim ilar rates to those under aerobic conditions in som e m arine sedim e nt (Caldwell et al. , 1998; Coates et al ., 1997). However, these high oil degradation rates under sulfate-reduc ing conditions have not been reported or dem onstrated in the field.

In summ ary, the potential of using oxygen am e ndm ent for enhancing oil biodegradation in
m arine shorelines and f r eshwater wetlands is lim ited. Tilling m ay be considered as an aeration strategy for enhancing oil biodegradation in the upper layer of sedim e nts in less sensitive habitats. Nitrate could be a potential alternativ e electron acceptor for use in a wide range of environm ents, but its use as an effective enhancer of biodegradation is questionable. More studies, particularly f i eld trials, are still required to evaluate the ef f e ctiveness of these strategies.

85Chapter 5 GUIDELINES FOR BIOREMED IATION OF MARINE SHORELINES AND FRESHWATER WETLANDS: DECISION-MAKING AND PLANNING Existing research and applications have dem ons trated that biorem ediation is an effective technology that can be used to treat certain oil- contam inated environm ents. Typically, it is used as a polishing step after conventional m echanical cleanup options have been applied, although it could also be used as a prim ary response strategy if the spilled oil does not exist as free product and if the contam inated area is rem o te enough not to require imm e diate cleanup or not accessible by m echanical tools. However, one of the m a jor challenges in the application of oil biorem ediation is lack of guidelines regarding the selection and use of this technology. Although extensive research has been conducted on oil biorem ediation in the last decade, m o st existing studies have been concentrated on either evalua ting the f easibility of biorem ediation f o r dealing with oil contam ination or testing favored pr oducts and m e thods (Mearns, 1997). Only a few lim ited operational guidelines for biorem ediation in m arine shorelines have been proposed (Lee, 1995; Lee and Merlin, 1999; Swannell et al ., 1996). The following two chapters will present a
m o re detailed and workable guidance for use by spill responders for the biorem ediation of
m arine shorelines and freshwater wetlands based on recent field studies and current understandings on biorem ediation processes. As a result of recent field studies (Lee et al , 1997b; Venosa et al ., 1996), we now know that there is usually little need to add hydrocarbon degradi ng m i croorganism s because this approach has been shown not to enhance oil degradation m o re than sim p le nutrient addition. Theref ore, this docum ent will only present guidelines f o r oil bi orem ediation using biostim ulation strategies,
m a inly nutrient addition. A general procedure or plan for the selection and application of biorem ediation technology is illustrated in Figure 5.1. The m a jor steps in a bi orem ediation selection and response plan include:
1. Pre-treatm e nt assessm ent – This step involve s the evaluation of whether biorem ediation is a viable option based on the type of oil that has been spilled, its concentration, the presence of hydrocarbon-degrading m i croorganism s, concentrat ions of background nutrients, the type of shoreline that has been im pacted, and other envi ronm ental factors (pH, tem p erature, presence of oxygen, rem o teness of the site, accessibility of the site and logistics, etc.).
2. Design of treatm e nt and m onito ring plan – After the decision is m a de to use biorem ediation, further assessm ents and planning are needed prio r to the application. This involves selection of

the rate-lim iting treatm e nt agents (e.g., nutrien ts), determ ination of application strategies for the rate-lim iting agents, and design of sam p ling and m onitoring plans.
3. Assessm ent and term ination of treatm e nt – Af ter the treatm e nt is im plem ented according to the plan, assessm ent of treatm e nt efficacy and determ ination of appropriate treatm e nt endpoints are perform e d based on chem ical, toxicological, and ecological analysis. This chapter covers operational guidelines for d ecision-m a king in the use of biorem ediation and describes the planning process for biorem ediati ng m arine shorelines and freshwater wetlands.

86The next chapter (Chapter 6) will present gui delines f o r assessm ent of f i eld results and establishm ent of appropriate treatm e nt endpoints.

St

e p 1: Pr etr e atm e n t As s e s s m e nt

St

e p 2 : Bior emedia tion Plan nin g

Step 3: A ssessme n t and Te rmi n at i o n Oil t

y pe an d
co n cen t ra t i o n Sh ore l i n e t y pe Bac k grou nd
nu t r ient conte n t

If

b i o r emed i a t i o n i s s e l e cted Clim a te , pri o r oi l
ex

posu r e and o t he r site c h ara c te risti c s

nu tri e n t produ c t s N u tri e n t ap pl i c at i on strategy Sa m p lin g an d
mo ni to ri ng pl an

T

oxicologi c a l an d e c o logi c a l
ana lysis An al y

s is o f

o il bi od e g

ra d a t i on a nd ph ys i c al lo ss

F i g ure 5.1 Pro c e d u r e s f o rth e s e le ctio n a n dap plic at ion of o i l spill biore m e d ia tion

875.1 Pre-treatment Assessment
Pretreatm e nt assessm ent involves som e pre lim inary investigations to assess whether biorem ediation is a viable option and determ ina tion of the rate determ ining process, which include the evaluation of (1) oil types and concentrations, (2) background nutrient content, (3) shoreline types, and (4) other environm ental factors such as the prevalent clim ate and prior oil exposures.
5.1.1 Oil type and concentration Oil Type As reviewed in Chapter 2, the biodegradability of different types of oils and petroleum products varies greatly depending on the distribution of oil com ponents. In general, the susceptibility to
m i crobial degradation for petroleum hydrocarbons is in the order of n-alkanes > branched alkanes > low-m o lecular-weight arom atics > cycl ic alkanes > high m o lecular-weight arom atics, although this pattern is not universal (Perry, 1984). The degradation rate for the sam e oil
com ponents m ay also vary significantly for differe nt oils. It has been found that the rate and extent of biodegradation of biodegradable co m ponents (e.g. n-alkanes) decreases with the increase of non-biodegradable fractions (e.g., resins and asphaltenes) (Uraizee et al ., 1998; W e stlake et al ., 1974). Therefore, the heavier crude oils are less likely biodegradable than lighter crude oils. McMillen et al. (1995) investigated the biodegrad ability of 17 crude oils with API gravity ranging from 14° to 45°. They concluded th at crude oil with greater than 30° API gravity can be considered readily biodegradable, and thos e with less than 20° API gravity (heavier oils) are slow to biodegrade. Sim ilar results were obtained by other researchers (Hoff et al ., 1995; Sugiura, et al ., 1997). W a ng and Bartha (1990) also inves tigated the effects of biorem ediation on residues of f

u el spills in soils. The results show ed that the treatability by biorem ediation f o r the fuel residues are in the order of jet fuel > heati ng oil > diesel oil. However, m o re work is still required to classify crude oils and refined products with respect to their theoretical am enability to cleanup by biorem ediation. The biodegradation potential of oils also depe nds on the weathering processes that alter oil
com positions and properties. For exam ple, evaporation leads to rem oval of the m o re toxic, lower
m o lecular weight com ponents from the spilled oil. Th erefore, there is less need to biorem ediate a spill of light petroleum products such as gasoline since it would evaporate rapidly. The form ation of water-in-oil em ulsion m ay increase m a ss tran sfer lim itation for oxygen and nutrients and decrease the oil biodegradation rate. Interactions be tween oil and various types of shorelines also play im portant roles in oil degradation, which will be discussed later. Field experience also suggested that oils that have been subjected to substantial biodegradation m i ght not be am enable to biorem ediation due to the accum u lation of polar com ponents in the oils (Bragg et al ., 1994; Oudet et al ., 1998).
Oil concentrations The concentration of oil is another im portant consideration in determ ining whether biorem ediation is a viable option. Very low c oncentrations of hydrocarbons in the environm ent m ay be inefficiently attacked by m i crobes (Foghtand W e stlake, 1987). For sites contam inated with oils at low concentrations, biodegradation is also less likely to be lim ited by nutrients or

88oxygen. Therefore, biorem ediation m ay not be e ffective in enhancing biodegradation in these cases. Natural attenuation m ay be a m o re viable option. High concentrations of hydrocarbons m ay cause inhi bition of biodegradation due to toxic effects, although the inhibitory concentration varies with oil com position. Therefore, there should be an optim um oil concentration range for biorem ediati on applications, below which degradation is not easily stim ulated, and above which inhibition o ccurs. However, this concentration range, particularly the m a xim um concentration of oil am enable to biorem ediation, has not been well quantified. Field experiences in Prince W illiam Sound, Alaska showed that less than 15 g oil/kg sedim e nts could be treated using biorem ediation (Swannell et al ., 1996). Xu et al . (2001) recently investigated the effect of oil conc entration in a m i crocosm study using weathered Alaska North Slope crude oil. The results showed that crude oil concentrations as high as 80 g oil/kg dry sand were still am enable to biodegr adation. Favorable oil concentrations for biorem ediation are also related to background condi tions, such as shoreline types, which will be discussed later in this chapter.
5.1.2 Background nutrient content
Since nutrient addition has been chosen as the prim ary strategy to enhance oil biodegradation,
assessm ent of background nutrient concentrations becom e s critical in determ ining whether biorem ediation is a viable option, whether natural attenuation should be considered, and/or which nutrient (nitrogen or phosphorus) should be added for oil biorem ediation.
As m e ntioned in Chapter 2, in m arine environm ents, nutrients are generally lim iting due to the naturally low nitrogen and phosphorus concentrat ions in seawater (Floodgate, 1984). Nutrient content is m o re variable in freshwater system s and is norm a lly abundant in freshwater wetlands (Cooney, 1984; Mitsch and Gosselink, 1993). Ho wever, background nutrients also depend on other site characteristics such as local industrial and dom estic effluents and agricultural runoff.
Recent field studies indicate that natural nutrient concentrations in som e m arine shorelines can be high enough to sustain rapid intrinsic rates of biodegradation without hum an intervention (Oudet et al ., 1998; Venosa et al. , 1996). The field trial in De laware showed that although biostim ulation with inorganic m i neral nutrients significantly accelerated the rate of hydrocarbon biodegradation, the increase in biodegradation rate over the intrinsic rate (i.e. slightly greater than twofold for the alkanes and 50% for th e PAHs) would not be high enough to warrant a
recomm e ndation to actively initiate a m a jor, perhap s costly, biorem ediation action in the event of a large crude oil spill in that area (Venosa, et al ., 1996). The high intrinsic biodegradation rate was attributed to the high bac kground nutrient concentrations (0.8 m g N/L on average) because the Fowler Beach area of Delaware Bay was adj acent to farm land. The relatively high organic content of both the Delaware Bay seawater and the underlying geology of the site and the presence of a saltwater m arsh several hundred m e ters landward from the beach could also account for the high nitrogen levels encounter ed. The study investigators observed that
m a intenance of a threshold nitrogen concentrati on of 3-6 m g N/L in the interstitial pore water was biostim ulatory for hydrocarbon biodegradation.

89A sim ilar conclusion was also reached in a field tr ial to evaluate the influence of a slow-release f e rtilizer on the biodegradation rate of crude oil spilled on interstitial sedim e nts of an estuarine environm ent in the bay of Brest, France (Oudet et al ., 1998). Due to the high background levels of N and P at the study site, no significant diffe rence in biodegradation rates was detected following nutrient addition. It was proposed that biorem ediation by nutrient enrichm e nt would be of lim ited use if background interstitial pore water le vels of N exceed 1.4 m g /L, which is close to the finding from the Delaware study (Venosa et al ., 1996). Phosphorus is another essential nutrient relate d to m i crobial growth. Although no field study of critical phosphorus concentrations on m arine s horeline and freshwater wetlands has been reported, it has been generally accepted that the optim al N:P ratio f o r m i crobial growth is in a range of 5:1 to 10:1. Therefore, the threshol d phosphorus concentration for m a intaining optim al hydrocarbon degradation can be derived based on th is ratio and critical nitrogen concentrations obtained from existing field studies. However, furthe r research is still required in determ ining the influence of phosphorus on oil biorem ediation unde r various m arine shoreline and freshwater wetland environm ents. These results suggest that, in the event of a catastrophic oil spill im pacting a shoreline, one of the first tasks in pretreatm e nt assessm ent is to m eas ure the natural nutrient concentrations within the interstitial water in that environm ent. If they ar e high enough, further investigation is required to determ ine whether such a nutrient loading is typi cal for that area and season (i.e., determ ine the im pact of chronic runoff from nearby agricultural practice and local industrial and dom estic effluents). The decision to use biorem ediation by addition of nutrients should be based on how high the natural levels are relative to the optim al or threshold nutrient concentrations.
5.1.3 Type of shorelines The characteristics or type of the contam inated shoreline also play an im portant role in the decision to use biorem ediation. This prelim inar y investigation involves the assessm ent of the need for biorem ediation based on wave and tida l energy, the sedim e nt characteristics, and geom orphology of the shoreline. Shoreline energy and hydrology Oil can be rem oved rather rapidly under high wa ve and tide influence, typically in rocky shorelines. In high-energy environm ents, biorem ediation products are also m o re difficult to apply successf ully since they can be washed out rapidly. High wave energy will also scour
m i croorganism s attached to the sedim e nt particle s, and dim i nish the net oil biodegradation rate that can be achieved. The Maine field study dem onstr ated that washout rate of nutrients from the biorem ediation zone will be strongly affected by th e wave activity of the contam inated beach. However, washout due to tidal activity alone is relatively slow, and nutrients will probably
rem a in in contact with oiled beach m a teri al long enough to effectively stim ulate oil biodegradation on low-energy beaches (Suida n and Wr enn, 2001; see Section 2.6.2). However, m any of the sam e characteristics that m a ke low-energy beaches favorable for biorem ediation cleanup from a nutrient persiste nce perspective m i ght m a ke other conditions unfavorable with respect to other im portant factor s. For exam ple, availability of oxygen is m o re

90favorable on high-energy beaches than on low-en ergy beaches. Aeration m echanism s for nearsurface coastal sedim e nts involve exchange of oxygenated surface water with oxygen-depleted pore water by wave-induced pum ping and tidal pum ping. For low energy beaches, tidal pum ping is the only likely aeration m echanism , and as a resu lt, the surface sedim e nts are m o re likely to be anoxic than are sim ilar depths on high-energy beaches (Brown and McLachlan, 1990). The probability of m o isture (or water activity) lim itation is also higher on low-energy beaches, because wave runup provides water to supratidal sedim e nts on high-energy beaches during neap tides (Suidan and Wr enn, 2001). Therefore, it is essential to thoroughly characterize the factors that are likely to be rate lim iting on each cont am inated site before deciding and designing a biorem ediation response strategy.
Shoreline substrate Although successful biorem ediation application and fi eld trials have been carried out on cobble,
m e dium sand, fine sand, and som e salt m arsh shorelines (Bragg et al ., 1994; Lee and Levy, 1991; Swannell et al. , 1999a; Venosa et al. , 1996), different shoreline substrates or sedim e nt types will affect the feasibility and strategies of using biorem ediation. In a 7-m onth field study, Lee and Levy (1991) com p ared the biorem ediati on of a waxy crude oil on a sandy beach and a salt m arsh shoreline. Terra Nova crude oil was a dded at two concentrations, 3% (v/v) and 0.3% (v/v) to beach sand and salt m arsh sedim e nts retained in in-situ enclosures in a low energy environm ent. The results showed that at the lo wer oil concentrations (0.3%) within the sand beach, oil biodegradation proceeded rapidly in both the fertilized plot and the unfertilized control. The application of a biorem ediation tr eatm e nt provided no advantage. However, at the higher oil concentrations (3%) on the sandy beach , oil biodegradation rates appeared to be nutrient lim ited and were enhanced by nutrient additi on. In contrast, the addition of nutrients to the salt m arsh sedim e nts containing the lower (0.3 %) oil concentration resulted in enhanced rates of biodegradation. This additional need for nutrients at the lower oil concentrations is consistent with the notion that nutrient dem ands within a salt m arsh environm ent are higher, due to the size of the m i crobial population within an organi c-carbon rich environm ent. At the higher oil concentration (3%) within the salt m a rsh sedim e nts, insignif i cant rates of oil degradation were reported following fertilization. The results clearly dem onstrated that the success of biorem ediation depends on the characteristics of the shoreline. On the sandy beach with low concentrations of oil, neither nutrient nor oxygen was a lim iting f actor. Under these conditions, nutrient enrichm e nt appears to provide little or no benef it, and
m onitored natural attenuation can be considered as an alternative. However, at higher oil levels, the m i crobial comm unity within the sand beach m ay becom e nutrient-lim ited, and biorem ediation treatm e nt could effectively enhance the rate of oil rem oval. In the salt m a rsh environm ent, nutrient addition was only effectiv e at low oil concentrations. Oxygen lim itation
was m o re likely at higher oil concentrations due to the finer particle size and higher organic content of the sedim e nt in these environm ents. Sim i lar results have been obtained in the field study conducted in a freshwat er wetland (Garcia-Blanco et al ., 2001b; Venosa et al. , 2002), which also indicated that oxygen availability was likely a m a jor rate-lim iting factor in the wetland environm ents. A field study sponsored by EPA and the Departm e nt of Fisheries and Oceans-Canada was recently conducted on the shorelin e of Nova Scotia to further investigate the potential of using biorem ediation in salt m arshes. Guidelines f o r oil biorem ediation in this type of shoreline will be available upon com p letion of the data analysis from this investigation.

915.1.4 Other factors
Prevalent climate Prevalent clim ate, the am bient tem p erature in particular, is an im portant consideration when assessing the f easibility of using biorem ediation. As discussed in Chapter 2, the am bient tem p erature of an environm ent affects both the properties of spilled oil and the activity or population of m i croorganism s. At low tem p eratures , the viscosity of the oil increases, delaying the onset of biodegradation (Atlas, 1981), and the volatility of toxic low-m o lecular-weight hydrocarbons is reduced. Although the rates of biodegradation generally decrease with decreasing tem p erature, biorem ediation has been te sted and applied successfully to enhance oil biodegradation in cold arctic, alpine, and Anta rctic environm ents (Margesin and Schineer, 1999). This is probably because psychrophilic bacteria are plentiful and generally the dom inant species in these m arine environm ents (Karrick, 1978).
A m o re im portant consideration regarding the effect of clim ate or weather on the use of biorem ediation perhaps is the seasonal factor. Si gnificant seasonal differences in the size of hydrocarbon degrader populations have been obs erved. The num bers of hydrocarbon degraders
m ay be m u ch lower during winter than summ e r in som e environm ents (Atlas, 1981). Oil biodegradation slows significantly and even ceases when the contam inated sedim e nts are frozen. Therefore, oil biorem ediation will be m o re eff ective during warm er seasons and probably should only be considered during the summ e r for cold environm ents such as arctic regions. Prior exposure to oil Prior exposure of a m i crobial comm unity to hydr ocarbons either from natural sources (e.g. chronic seeps and plant derived hydrocarbons) or as a result of pollution (e.g. spills and waste disposal) m a y affect the rate at which subse quent hydrocarbon input can be biodegraded (Leahy and Colwell, 1990). Those environm ents with a hi story of oil pollution or natural oil inputs have been found to have a m u ch higher percentage of hydrocarbon degraders and a generally greater potential of hydrocarbon degradation than prev iously unpolluted areas (Atlas, 1981; Lee and Levy, 1987, Pritchard and Costa, 1991). Therefore, for oil biorem ediation in environm ents with no prior oil exposure, there m a y be a lag and adaptation period before any significant oil biodegradation occurs. This usually is not a con cern since biorem ediation itself is a relative slow process and typically is used as a polishi ng step after conventional m echanical cleanup operations. In contrast, those environm ents withprior exposure to oil need a shorter lag period bef o re initiation of biodegradation and thus will likely have a higher potential f o r oil biodegradation. Thus, this type of environm ent is generally considered a favorable condition for using biorem ediation.
5.1.5 Summary of pretreatment assessment
In summ ary, the following pretreatm e nt assessm en ts should be conducted to determ ine whether biorem ediation is a viable option in response to a spill incident:

• Determ ine whether the spilled oil is potential ly biodegradable – Light petroleum products and light crude oils (API gravity > 30°) are re latively biodegradable; products rich in norm a l

92alkanes are relatively biodegradable; heavy crude oils (API gravity < 20°) and residual fuel oils, which are high in polar com pounds (as phaltenes and resins) are less biodegradable. High concentrations of oil m ay also inhibit biodegradation.

• Determ ine whether the nutrient content at the im pacted area is likely to be an im portant lim iting factor by m easuring the background nutri ent concentrations within the interstitial water in that environm ent – The decision to use biorem ediation by addition of nutrients should be based on how high the natural levels are relative to the optim al or threshold nutrient concentrations (e.g., > 2 m gN/L on sa ndy m arine shorelines). It should also be determ ined if the natural nutrient concentrations present are typical of th e area or sporadic. If sporadic, biostim ulation m ay still be appropriate when the nutrient levels f a ll to lim iting values; if chronic, biostim u lation m ay not be necessary.

• Determ ine whether the shoreline characteristic s are favorable for using biorem ediation – High-energy rocky beaches and som e low ener gy shorelines such as som e wetlands are considered not likely to be very am enable to nutrient addition.

• Determ ine whether clim atic or seasonal conditions are favorable for using biorem ediation – biorem ediation will be m o re ef f ective duri ng warm er seasons, particularly in cold environm ents. Prior exposure to oil will also be a favorable but not a solely determ inative condition for selecting biorem ediation.
5.2 Selection of Nutrient Products
After biorem ediation is determ ined to be a potentially effective cleanup option based on the prelim inary investigations, f urther assessm ents and planning are needed before its application. The first task is to select appropriate nutrient products through both screening tests and
assessm ents based on characteristics of the contam inated site.
5.2.1 Nutrient selection based on efficacy and toxicity To assist response personnel in the selection and use of spill biorem ediation agents, it is useful to have som e sim p le, standard m e thods for screening the perform ance and toxicity of biorem ediation products as they becom e available (Blenkinsopp, et al ., 1995; Haines et al ., 1999, Lepo and Cripe, 1998a). One of the m o st com p rehens ive exam ples of such protocols is the tiered approach developed by EPA, in cooperation with the National Environm ental Technology Applications Center (NETAC, 1993; Thom as et al ., 1995). Conducting treatability tests using
m i cro- or m e socosm s is another comm only used approach. EPA/NETAC protocols The NETAC/EPA protocols consist of five progre ssive tiers, which increase in environm ental cost and com p lexity with each tier of testing (Tab le 5.1). The approach begins with a Base Tier in which basic inf o rm ation on the agent’s toxi city is gathered based on a review of its form ulation. During this tier, the presence of chem icals or biological com ponents that are
norm a lly considered unacceptable (i.e. pathogens , carcinogens, or hazardous substances) would

93be identified. Tier I provides the basis for a pre lim inary evaluation of whether an agent could be effective and safely applied, which includes a de scription of how the product will be used, and inform ation on previous usage. Tier II provides em pirical evidence through the use of laboratory shake flask treatability studies to estim a te a pr oduct’s effectiveness. This tier also provides inf o rm ation on the relative changes in aliphatic and arom atic oil constituent concentrations over tim e and the total hydrocarbon degrading m i crobial activity. Tier III proposes the use of flowthrough m i crocosm system s to study biodegradation e ffectiveness. Tier IV is the use of field dem onstrations to predict a product’s potential e ffectiveness in the natural environm ent. Tiers III and IV are no longer considered viable options when evaluating a product for use in an oil spill due to overwhelm ing econom ic considerations. It is clear that field studies can provide the m o stconvincing dem onstration of the effectiveness of oil biorem ediation because laboratory studies sim p ly cannot sim u late real world conditions such as spatial heterogeneity, clim ate change, and m a ss transfer lim itations. Since conducting a field study just to determ ine that a product m i ght wo rk is unrealistic and econom ically burdensom e, the practical approach in selection of nutrient products for the biorem ediation of an oil spill would be through laboratory tests, m i crocosm tests in particular, as well as evaluations based on existing field study results in sim ilar environm ental conditions.
Table 5.1 Biorem ediation product test protocol s developed by EPA and in cooperation with NETAC. Test Levels Description
Base Tier

Tier I

Tier II

Tier III

Tier IV Collection and analysis of basic inform ation on product safety including form ulation and acceptabilityof its chem ical or biological
com ponents Feasibility assessm ent concerning production capabilities, potential ef

f e ctiveness, and saf e ty certif ication , including a description of how the product will be used, and inform ation on previous usage Efficacy and safety evaluation usi ng shake flasks to com p are the degradation of artificially-weathered crude oil in natural seawater with and without a biorem ediation agent Efficacy and safety evaluation usi ng m i crocosm system s to sim u late various environm ents (e.g. open water, beaches, and m arshes) Efficacy and safety evaluation through a field dem onstration

Nutrient selection through microcosm tests The laboratory treatability tests, especially well -designed m i crocosm tests, are m o st comm only used approaches to determ ine the type and level of am endm ents, such as the types of f

e rtilizer

94and the optim al nutrient concentrations. A good exam ple of these m i crocosm s is illustrated in Figure 5.2, which has been used for various screening and treatability tests (Ahn, 1999, Du et al ., 1999, Xu et al ., 2001). These m i crocosm s have at least th ree advantages over the batch reactors that are often used in this type of study: (1) they are connected to respirom eters, allowing nondestructive acquisition of kinetic data by con tinuously recording the oxygen consum ption that
accom p anies oil biodegradation; (2) they are open system s that can sim u late the nutrient washout that will occur in contam inated intertidal zone s; and (3) they are designed to sim u late tidal flushing by filling and draining on a 12-hour cycle, thus sim u lating the periodic anoxia that can occur due to tidal flooding. The potential for oxygen lim i tation in these reactors is a particularly im portant advantage over m o re conventional m i cr ocosm s
. Because nutrient concentrations do not lim it the oil biodegradation rate when oxygen becom e s sufficiently depleted, these system s will provide m o re realistic estim ates of m a xim um biodegradation rates than well-aerated shake flasks will provide. Also, som e fertilizers contai n large am ounts of readily biodegradable organic
com pounds that can accentuate dissolved oxygen depletion. Although they m i ght be very effective in well-aerated m i crocosm s , oxygen availa bility can lim it their eff ectiveness in the field (Lee et al ., 1995a).

Co nn ec t e d t o Re sp i r o m e t e r
Gl a s s Be ad s

Oi l Co nt a m i n a t ed Sa n d s A B

Ti me r

Mi n e r a l S a l t s So l u t i o n
Pu m p S a m p ling Po r t Figure 5.2 Schem a tic diagram of a beach m i croc osm for laboratory treatability testing of oil spill biorem ediation treatm e nt.

95Using this m i crocosm system , Xu et al . (2001) investigated the e ffect of different nitrogen nutrients on the biorem ediation of weathered Alas ka North Slope crude oil under sim u lated tidal conditions. Three oil concentrations of 5, 20, and 80 g oil/kg dry sand were used. Two types of nitrogen nutrients (KNO 3 and NH 4 Cl) were applied at a con centration of 100 m g -N/L. Oil biodegradation was evaluated by m onitoring CO 2 production, oxygen uptake, nitrogen
consum ption, as well as oil constituent analysis. Results indicated that m o re biom ass growth occurred in the subm erged (som etim es anoxic) portion of the sand, and better oil degradation was observed in m i crocosm s to which nitrate-ni trogen was applied. This result suggested that nitrate m i ght also have enhanced oil biorem ediation by serving as an electron acceptor when oxygen was lim iting. However, the role of nitrate still requires further investigation. Ram s tad and Sveum (1995) also found that nitrate had th e m o st pronounced effect in stim ulating oil degradation when com p aring the effect of nitr ate, amm onia, and an organic nitrogen product on biodegradation of topped Statfjord crude oil ina continuous-flow seawater columnsystem

. Effect of nutrient type m ay also depend on the properties of shoreline substrates. Jackson and Pardue (1999) found that addition of amm onia as com p ared to nitrate appeared to m o re effectively sim u late degradation of crude oil in salt m arsh soils in a m i crocosm study. The
amm onia requirem e nt was only 20% of the concentr ation of nitrate to achieve the sam e increase of degradation. The authors concluded that amm onia was less likely to be lost from the
m i crocosm s by washout due to its higher adso rptive capacity to sedim e nt organic m a tter. However, in a m i crocosm study using sandy sedi m e nts, Suidan and Wr enn (2000) found that there were no significant differences in the nutri ent washout rates or the abilities of amm onium and nitrate to support oil biodegradation. Thes e results suggest that although cation-exchange adsorption m a y be an im portant difference betw een amm onium and nitrate in sedim e nts with high cation-exchange capacities (CECs), such as m a rsh sedim e nts, it is unlikely to be significant in sedim e nts with low CECs, such as sand. Toxicity and other environmental impacts In addition to dem onstrating the efficacy of nutri ent products in enhancing oil degradation, it is also necessary to dem onstrate that biorem edia tion products have low toxicity and do not produce any undesired environm ental and ecological effects. Various toxicity test protocols have been developed (NETAC, 1993; Lepo and Cripe, 1998a, S ee Section, 3.5). For exam ple, the EPA Tier II safety evaluation consists of 7-day toxi city tests with the biorem ediation product (without oil) in natural seawater using a crustacean ( Mysidopsis bahia , my sid), and a fish ( Menidia beryllina , the inland silverside). Additional Tier II toxicity tests evaluate the potential f o r interaction between the product and the water-solubl e-fraction of a weathere d crude oil. Indirect effects of nutrient products should also be ev aluated, which include oxygen depletion through increase in organic carbon or eutrophication, and enhanced production of toxic oil degradation
m e tabolites (Lepo and Cripe, 1998b). So far, no detrim ental effects from biorem ediation by nutrient enrichm e nt have been observed following actual field operations (Prince, 1993; Mearns et al ., 1997). However, the possibility that harm ful effects m i ght occur rem a ins. For exam ple, oxygen depletion and production of
amm onia from excessive applications of a fish -bone m eal fertilizer during a field study caused detrim ental effects, including a slowing in oil degradation rates and toxicity reduction rates

96m easured by Microtox  Solid-Phase Test (Lee et al ., 1995b). For safety reasons, proper ecotoxicological assessm ent is always necessary in selecting nutrient products.
5.2.2 Environmental factors affecting nutrient selection Nutrient selection also depends on environm ental factors such as tem p erature, shoreline energy , and substrate. A field study conducted by Lee et al . (1993) indicated that the effectiveness of specific nutrient form ulations m i ght be in fluenced by tem p erature conditions. The study investigated the efficacy of water-soluble inor ganic fertilizers (amm onium nitrate and triple superphosphate) and a slow release fertilizer (s ulfur-coated urea) to enhance the biodegradation of a waxy crude oil in a low ener gy shoreline environm ent. The results showed that at tem p erate conditions above 15°C, the slow-release fertilizer appeared to be m o re effective in retaining elevated nutrient concentrations within the se dim e nts and m o re effective in enhancing oil degradation than water-soluble fertilizers. Howe ver, lower tem p eratures were found to reduce the perm eability of the coating on the slow-release fertilizer and suppressed nutrient release rates. W a ter-soluble fertilizers such as amm onium nitrate were then recomm e nded under these tem p erature conditions. The action of wave, tide, and sedi m e nt type will also affect the selection of nutrients. Som e studies suggested that oleophilic fertilizers m i ghtbe m o re suitable for use in high-energy and coarse-grained beaches (Sveum et al ., 1994; Sveum and Ladousse, 1989; See Chapter 4), although stronger evidence is needed to confir m

this suggestion. Therefore, for optim al effectiveness, the nutrient selection should al ways take into account the environm ental conditions, the type of contam inated shoreline, and the m e thods of application, which will be discussed later in this chapter (Lee et al ., 1993; Prince, 1993; Swannell et al ., 1996).
5.3 Determination of the Optimal Nutrient Loading and Application Strategy After the initial selection of nutrient products that m e et the requirem e nts of efficacy and safety, the next step is to determ ine the proper nutri ent loading and the best nutrient application strategies. Major considerations in this task include the determ ination of optim al nutrient concentration, frequency of addition, and m e thods of addition. Finally, selection of appropriate nutrient products should also be conducted in conjunction with this process.
5.3.1 Concentration of nutrients needed for optimal biostimulation Since oil biodegradation largely takes place at the interface between oil and water, the ef

f ectiveness of biostim ulation depends on the nutrient concentration in the interstitial pore water of oily sedim e nts (Bragg et al ., 1994; Venosa et al ., 1996). The nutrient concentration should be
m a intained at a high enough level to support m a xi m um oil biodegradation based on the kinetics of nutrient consum ption. Higher concentrations will not only provide no a dded benefit but also
m ay lead to potentially detrim ental ecological and toxicological im pacts. Studies on optim al nutrient concentrations have been conducted both in the laboratory and in the field. Boufadel et al ., (1999a) investigated the optim al nitrate concentration for alkane biodegradation in continuous flow beach m i crocos m s (Figure 5.2) using heptadecane as a m odel

97alkane imm obilized onto sand particles at a loading of 2g heptadecane/kg sand. They determ ined that a continuous supply of approxim a tely2.5 m g N/L supported m a xim um heptadecane biodegradation rates. Du et al. (1999) also investigated the optim al nitrogen concentration for oil biodegradation using weathered Alaska North Slope crude oil in the sam e m i crocosm s with an oil loading of 5g/kg sand. The results showed that nitrate concentrations below approxim a tely 10
m g N/L lim ited the rate of oil biodegradation. Th e higher nutrient requirem e nt was attributed to the m o re com p lex substrate (crude oil) co m p ared to the pure heptadecane of Boufadel et al. (1999a). The m o re com p lex substrate (crude oil) of Du et al. (1999) also likely selected a different population of degraders than those that grew on the pure heptadecane (Boufadel et al. , 1999a), which m i ght have contributed to the diffe rent growth rate characteristics observed. Ahn (1999) further studied the eff ect of nitrate concentrations under tidal flow conditions instead of continuous flow. He used the sam e beach m i crocosm s as Du et al. (1999) f illed with sand loaded with weathered Alaska North Slope cr ude oil at 5g/kg sand. A nutrient solution with nitrate concentrations ranging from 6.25 to 400 m g N/L was supplied sem i -diurnally to sim u late tidal flow. The results indicated that the opt im um nitrate concentration for m a xim um oil biodegradation rate was over 25 m g N/L. Som e labor atory studies have reported that greater than 100 m g N/L was required to stim ulate m a xim um biodegradation rates (Atlas and Bartha, 1992; Reisf e ld et al. , 1972), but this observation probably reflects a stoichiom e tric rather than a kinetic requirem e nt, since these experim e nts were conducted in closed batch reactors.
Com p ared to the results from laboratory studies , nutrient concentrations that supported high oil biodegradation rates were found to be lower in fi eld studies. For exam ple, the field tests that were conducted after the Exxon Valdez oil spill in Prince W i lliam S ound, Alaska showed that the rate of oil biodegradation was accelerated by averag e interstitial nitrogen concentrations of about 1.5 m g N/L (Bragg et al ., 1994). A sim i lar result was obtaine d from the study conducted in the Bay of Brest, France, in which nitrogen was not a lim i ting f a ctor when the interstitial pore water concentrations exceeded 1.4 m g N/L (Oudet et al ., 1998). The Delaware field trial also showed that the background nitrate concentration (0.8 m g N/L) was sufficient to support fairly rapid natural (but not m a xim a l) rates of alkane and PAH biodegradation (Venosa et al ., 1996). Increasing the average nitrate concentration in th e biorem ediation zone of the experim e ntal plots to between 3 and 6 m g N/L resulted in a m oderate increase in the oil biodegradation rate.
Observations from the referenced field studies s uggest that concentrations of approxim a tely 1 to 2 m g /L of available nitrogen in the interstitia l pore water is suf f i cient to m eet the m i nim um nutrient requirem e nt of the oil degrading m i cr oorganism s for the approxim a tely 6-hour exposure tim e to the contam inated substrate during a tidal cycle. However, laboratory m i crocosm results as well as the Delaware field study suggest that higher concentrations of nitrogen can lead to accelerated hydrocarbon biodegradation rates. Sin ce the m i nim um nitrogen concentration needed to satisf ythe nitrogen dem and in a tidal cycle is 1 to 2 m g N/L, and si nce concentrations of nitrogen in pore water that lead closer to m a xim um rates of biodegradation can be several-fold to
as m u ch as an order of m a gnitude higher, it is recomm e nded that biostim ulation of oil im pacted beaches should occur when nitrogen concentrations of at least 2 to as m u ch as to 5-10 m g N/L
are m a intained in the pore water with the deci sion on higher concentrations to be based on a broader analysis of cost, environm ental im pact, and practicality. In practice, a saf e ty f a ctor should be used to achieve target concentrat ions, which will depend on anticipated nutrient

98washout rates, selected nutrient types, and app lication m e thods. For exam ple, in the Delaware study, since nitrate in the interstitial pore wa ter was quickly diluted to background levels whenever the incom i ng tide com p letely subm er ged the plots, water-soluble nutrients were applied every day using a sprinkler system . A 100- fold safety factor to account for dilution was used to achieve the 3-6 m g /L average interstitial pore water concentrations experienced at Delaware. A lower safety factor m ay be needed when using slow release nutrients.
5.3.2 Nutrient application strategies Once the optim al nutrient concentrations have b een determ ined, the next task is to design nutrient application strategies, which include nutrient application frequency and delivery
m e thods. Frequency of nutrient addition The f r equency of nutrient addition to m a intain the optim al concentration in the interstitial pore water m a inly depends on shoreline types or nutrient washout rates. On m arine shorelines, contam ination of coastal areas by oil from offshore spills usually occurs in the intertidal zone where the washout of dissolved nutrients can be extrem ely rapid. Oleophilic and slow-release form ulations have been developed to m a intain nutri ents in contact with the oil, but m o st of these rely on dissolution of the nutrients into the aqueous phase before they can be used by hydrocarbon degraders (Safferm an, 1991). Therefor e, understanding the transport of dissolved
com pounds in intertidal environm ents is criti cal in designing nutrient addition strategies, no
m a tter what type of f

e rtilizer is used. The Maine field study on nutrient hydrodynam i cs (See Section 2.6) has dem onstrated that during spring tide, nutrients can be com p letely rem ove d from a high-energy beach within a single tidal cycle. But it m ay take m o re than two weeks to achieve the sam e degree of washout froma lowenergy beach. W a shout during the neap tide can be m u ch slower because the biorem ediation zone will be only partially covered by water in th is period. Since nutrients m ay be com p letely washed out from high-energy beaches within a fe w days, and rem a in in low energy beaches for several weeks, the optim al frequency of nutrient application should be based on observations of the prevalent tidal and wave conditions in the bi orem ediation zone. For exam ple, a daily nutrient application m ay be needed for a high-energy b each during spring tide. But weekly or m onthly additions m ay be sufficient for low-energy beaches when the nutrients are applied during neap tide. Nutrient sam p ling, particularly in beach por e water, m u st also be coordinated with nutrient application to ensure that the nutrients becom e distributed throughout the contam inated area and that target concentrations are being achieved. The frequency of nutrient addition should be adjusted based on the nutrient m onitoring results. Methods of nutrient addition Nutrient application m e thods should be dete rm ined based on the characteristics of the contam inated environm ent, physical nature of the selected nutrients, and the cost of the application. Shoreline energy and geom etry are im portant fact ors in determ ination of nutrient application
m e thods. The study in Maine suggested that surface application of nutrients m a y be ineffective

99on high-energy beaches because wave action in the upper intertidal zone m ay cause nutrients from the surface layers of the beach to be diluted directly into the water column , resulting in their imm e diate loss from the biorem ediation zone. Da ily application of water-soluble nutrients onto the beach surface at low tide could be a feasible approach (Venosa et al ., 1996), although this
m e thod is highly labor-intensive. Nutrients that are released from slow-release or oleophilic f o rm ulations will probably behave sim ilarly to water-soluble nutrients with respect to nutrient washout. Form ulations with good long-term rel ease characteristics probably will never achieve optim al nutrient concentrations in environm ents w ith high washout rates. Therefore, they will not be effective on high-energy beaches unless the re lease rate is designed to be high enough to achieve adequate nutrient concentrations while the tide is out. Another potentially ef f ective strategy is the subs urface application of nutrients onto high-energy beaches. W i se et al. (1994) found that application of nutri ents through a trench or subsurface drain placed above the high-tide level, rather th an directly on the beach by sprinklers, would result in significantly longer retention tim es. However, since nutrients m ove downward and seaward during transport through the intertidal zone of sandy beaches, nutrient application strategies that rely on subsurface introduction m u st provide som e m echanism for insuring that the nutrients reach the oil-contam inated area ne ar the surface. The approach that was proposed by W i se et al . (1994) assum e s that nutrients dissolved in the freshwater plum e will be brought into contact with the oiled beach m a terial peri odically by the rising tide because the freshwater plum e should float on top of the saltwater. However, the finding of freshwater trapped between two saltwater wedges by Boufadel et al . (1999) indicates that subsur face injection of nutrients above the high tide level would likely not be an effective m e thod for providing nutrients to the biorem ediation zone.
Com p ared to high-energy shorelines, application of nutrients on low-energy beaches is m u ch less problem atic. Since washout due to tidal activity alone is relatively slow (Suidan and Wr enn, 2001), surface application of nutrients is an e ffective and econom ical biorem ediation strategy on low-energy beaches. The physical form s of fertilizers are also critical in determ ination of appropriate nutrient application m e thods. Generally, available fertilizers canbe classified into four types in term s of their physical form s: (1) slow release fertilizer briquettes, (2) dry granular fertilizers, (3) liquid oleophilic f e rtilizers, and (4) water-soluble inorganic nutrient solutions (Glaser et al ., 1991; Swannell et al ., 1996). The application of the briquette form s is problem atic in regards to buoyancy of the briquettes and redistri bution by tide and wave action (Glaser et al., 1991). The
m e thod used during the Exxon Valdez spill involved packing the briquettes in m e sh bags tethered to steel bars driven into the beach subsurface. The poor distribution problem occurs by channeling of nutrients vertically down the beach rather than lateral spreading. Dry granular f e rtilizers can be slow-release (e.g., Custom blen in Alaska) or water-soluble, solid granules (e.g., prilled amm onium nitrate). Granularf e rtilizers are easier and m o re f l exible to apply using comm e rcially available whirlybird -type hand spreaders. Although this type of fertilizer is also subject to washout by wave and tidal action, dry granular fertilizers are probably the m o st cost-effective way to control nutrient concentrations. Liquid oleophilic nutrients, such as Inipol EAP 22, are also relatively easy to a pply by using hand-held or backpack sprayers. One

100of the problem s when usi ng Inipol EAP 22 during the Exxon Valdez spill cleanup was its low pour point (11°C), which led to clogging of spraying nozzles and poor unifo rm ity of application (Glaser et al., 1991). The problem was later resolved by warm ing the product. This type of f e rtilizer is signif i cantly m o re expensive than granular f e rtilizers. W a ter-soluble nutrient solutions are norm a lly delivered to the beach by a sprinkler system after dissolving nutrient salts in a local water source. Although this type of nutrient m ay be easier to m anipulate to m a intain target concentrations in inters titial pore water, its application m a y require m o re com p licated equipm ent such as large m i xing tanks, pum ps , and sprinklers. Based on current experiences and understandings, application of dry granular fertilizer to the im pact zone at low tide is probably the m o st cost-ef f ective way to control nutrient concentrations.
5.4 Sampling and Monitoring Plan To properly evaluate the progress of biorem edia tion, a com p rehensive and statistically valid
sam p ling and m onitoring plan should be developed before the application of biorem ediation. The
sam p ling and m onitoring plan should include im portant efficiency and toxicity variables, environm ental conditions, and sam p ling strategies.
5.4.1 Important variables
Im portant variables to be m onitored in an oil biorem ediation project include lim iting factors for oil biodegradation (e.g., interstitial nutrient and oxygen concentrations), evidence of oil biodegradation (e.g., concentrations of oil and its com ponents), m i crobial activity (e.g., bacterial
num bers), environm ental effects (e.g., ecotoxic ity levels) and other environm ental conditions (e.g., tem p erature and pH). A com p rehensive m onitoring plan proposed for a biorem ediation field study is listed in Table 5. 2. A m onitoring plan for a full-scale biorem ediation application should be sim ilar to this and at least should include those critical m easurem ents. Since oil biodegradation in the f i eld is usually lim i ted by availability of nutrients, nutrient analysis, particularly the nutrient concentrations in the pore water, is one of the m o st im portant m easurem ents in developing proper nutrient addition strategies and assessing the effect of oil biorem ediation. The frequency of nutrient sam p ling m u st be coordinated with nutrient application, m a king certain that the treatm e nt is reaching and penetrating the im pact zone, target concentrations of nutrients are being achieved, and toxic nutrient levels are not being reached. Otherwise, nutrient application strategies shoul d be adjusted accordingly. The location from which nutrient sam p les are collected is also im portant. Recent research on solute transport in the intertidal zone has shown that nutrients can rem a in in the beach subsurface for m u ch longer tim e periods than in the biorem ediation zone (Wrenn et al ., 1997a). Nutrient concentration profiles along the depth of the oil-contam inated region can be m onitored by using m u lti-port sam p le wells or sand sam p les collected from the oil-contam inated region.
The sam p ling depth should be determ ined based on th e prelim inary survey of oil distribution. It can be established by determ ining the m a xim um de pth of oil penetration, then adding a safety f actor, which will be chosen based on the observed variation in oiled depth, to ensure that the

101sam p les will encom p ass the entire oiled depth thr oughout the project. The safety factor will be
m odified if observations during the biorem ediati on application suggest that the depth of oil penetration has changed.
The success of oil biorem ediation will be judged by its ability to reduce the concentration and environm ental im pact of oil in the field. As discussed in Chapter 3, to effectively m onitor biodegradation under highly heterogeneous conditions , it is necessary that concentrations of specific analytes (i.e., target alkanes and PAHs) within the oil be m easured using chrom a tographic techniques (e.g., GC/MS) and ar e reported relative to a conservative biom arker such as hopane. On the other hand, from an opera tional perspective, m o re rapid and less costly analytical procedures are also needed to satis fy regulators and responders on a m o re real tim e, continual basis. Existing TPH technologies are ge nerally not reliable and have little biological significance (See Chapter 3). TLC-FID seem s to be a prom ising screening tool for m onitoring oil biodegradation (Stephens et al ., 1999).
Table 5.2 Monitoring plan for a biorem ediation field study Recomm e nded Methods Analysis Matrix *dissolved nitrogen Sedim e nt (interstitial pore water) extract in acidified 0.1% NaCl 4500-N or g D (persulfate oxidation to NO 3 - ) 4500-NO 3 - F (autom ated Cd-reduction) dissolved phosphorus Sedim e nt (interstitial pore water) extract in acidified 0.1% NaCl 4500-P B.5 (persulfate oxidation) 4500-P E (ascorbic acid m e thod) *residual oil Sedim e nt extract into m e thylene chloride analyze com ponents by GC/MS-SIM dissolved oxygen Aqueous Hach ® high range assay pore-water pH Aqueous potentiom etric with com b ination electrode

m i crobial populations Sedim e nt MPN for alkane and PAH degraders

m e tabolic activity Sedim e nt CO 2 production from sand slurries *toxicity of residual oil Sedim e nt Microtox ® Solid-Phase Test toxicity of residual oil Sedim e nt 10-day am phipod survival bioassay toxicity of pore-water Aqueous Microtox ® Acute Test toxicity of pore-water Aqueous sea-urchin f e rtilization bioassay bioaccum u lation potential SPMDs ** 2-week adsorption into SPMDs beach profile surveying using fixed m arkers (e.g., wells, plot boundary m arkers) in intertidal zone beach profile surveying relative to fixed benchm arks in the supratidal zone * Critical m easurem ents ** Sem i -perm eable m e m b rane devices

102In addition to m onitoring the treatm e nt effi cacy for oil degradation, the biorem ediation
m onitoring plan should also incorporate re liable ecotoxicological endpoints to docum ent treatm e nt effectiveness for toxicity reduc tion. Comm only used ecotoxicity m onitoring techniques, such as the Microtox ® assay and an invertebrate survival bioassay, have been
summ arized in chapter 3. These m i croscale bioassays m ay provide an operational endpoint indicator f o r biorem ediation activities on the basis of toxicity reduction (Lee et al ., 1995b). Considerations for selecting an appropriate biorem ediation endpoint based on both oil degradation and toxicity reduction will be presented in detail in the next chapter.
Other im portant variables in a com p rehens ive m onitoring plan include site background conditions (e.g., oxygen, pH and tem p erature) and shoreline profiles. Oxygen availability is crucial for rapid biorem ediation, because hydr ocarbon biodegradation is prim arily an aerobic process. Therefore, dissolved oxygen (DO) in the pore water should be m onitored on a regular basis. The frequency of DO sam p ling should also be coordinated with nutrient application, particularly when organic nutrients are used (Lee et al ., 1995b; Sveum and Ram s tad, 1995; See Section 4.1.3), to insure that anoxic conditions do not result. Wh en oxygen does becom e lim ited, the nutrient dosage and application frequency s hould be adjusted accordingly. Alternatively, oxygen am endm ent strategies, such as tilling or addition of chem ical oxygen sources, m ay be considered, although these approaches are likely to be expensive and potentially environm entally hazardous (see Chapter 4).
Measurem ent of pH in the pore water is also im portant in m onitoring oil biorem ediation. Biodegradation of oil in m arine environm ents is optim al at a pH of about 8 (Atlas and Bartha, 1992). The pH of seawater is usually around 8. 5, which is adequate to support rapid oil biodegradation.
5.4.2 Statistical considerations in the sampling plan
To ensure that m onitored results reflect the real ity in a highly heterogeneous environm ent, it is im portant that a biorem ediation sam p ling planbe designed according to valid statistical principles such as random ization, replication, and proper control.
To m i nim i ze bias, a random sam p ling plan should be used to evaluate treatm e nt effects and their variance within the biorem ediation zone. Fo r sam p les with a high degree of spatial heterogeneity, which will be the case for m o st oi l spill sites, stratified sam p ling strategies should be used. For exam ple, the sam p ling field on a m arine shoreline can be divided into a num ber of sectors or quadrants based on the hom ogeneity of geom orphology within each sector (e.g., upper and lower intertidal zones), and independent sam p les should be taken in each sector according to the rule of proportionality (e.g., taking m o re sam p les in m o re heavily oiled sites). Although econom ic factors m ay be restrictive, effort s should be m a de to ensure that an adequate
num ber of sam p les are taken to reach a given accu racy and confidence. Power analysis should be used to assist in the determ ination of sam p le replications required in a m onitoring plan. A statistical power test was perform e d in the De laware field study to determ ine the num ber of replicates that would be needed in future studi es to detect significant treatm ent differences under

103sim i lar conditions (Venosa et al ., 1996). The study indicated that the required replicates to detect treatm e nt effects depend on expected variance and expected treatm e nt differences. For exam ple, if oil distribution and shoreline characteristics are highly heterogeneous, variance will be high, thus requiring m o re replicates to detect signifi cant treatm ent effects. If background nutrients are high, treatm e nt differences will be low, and m o re replicates will also be required. By com p aring three shoreline assessm ent designs used for the Exxon Valdez oil spill, Gilf illan et al . (1999) also proposed several strategies to increase power (i .e., the probability that significant differences between two or m o re treatm e nts are detected when indeed they exist). One of the approaches to increase power is to select sam p ling sites from only the m o st heavily oiled locations. This strategy m ay not be feasible for assessing the o il degradation within the whole biorem ediation zone, although it m ay be useful for evaluating the effect of biorem ediation on ecological recovery since the ecological injury m o st likely occurs at the heavily oiled locations.
A control area norm a lly refers to a set-aside untreated site, which has sim ilar physical and biological conditions as the treated site. Although on-scene coordinators prefer not to leave oiled sites uncleaned, it is difficult to assess the true im pact of a treatm e nt without control or set-aside areas (Hoff and Shigenaka, 1999). Wh en selecting control areas, one m u st consider not only the sim i larity of the conditions but also the ef f ect of sand and nutrient exchanges between the treated and untreated areas (See Section 6.1.2).
5.5 Considerations for Freshw ater Wetland Bioremediation Guidelines proposed in earlier sections are m o stlyderived from studies and practices on m arine shorelines. However, freshwater conditions or habitats m ay differ sufficiently from m a rine situations so that a sim p le transfer of res ponse strategies m ay not be necessarily the m o st appropriate. Special considerations for oil biorem ediation in freshwater wetlands are summ arized here based on current understandings, particularly the findings of the St. Lawrence River field study (Garcia-Blanco et al ., 2001b; Venosa et al., 2002; Lee et al ., 2001a).
5.5.1 Characteristics of freshw ater w e tlands
W e tlands occupy the interface between terrestrial and aquatic system s. They have been defined by EPA and the U.S. ArmyCorps of Engineers as: “ Those areas that are inundated or saturated by surface or groundwater at a frequency or duration to support, and that under normal conditions do support, a prevalence of vegetati on typically adapted for life in saturated soil conditions. Wetlands generally include swamps, marshes, bogs, and similar areas .” (EPA, 40 CFR 230.3 and Corps, 33 CFR 328.3). W e tlands norm a lly should have the following three characteristics or diagnostic param e ters: (1) at l east periodically, the land supports predom inantly hydrophytes (i.e., plants adapted to the flooded conditions), (2) the substrate is predom inantly undrained hydric soil (i.e., a soil with unique physical and chem ical characteristics, such as highly reduced conditions, due to repeated and prolonged saturation), and (3) the substrate is saturated with water or covered by shallow wate r for a significant part of the growing season each year (Greene, 2000; Mitsch and Gosselink, 1993).
W e tland ecosystem s have enorm ous ecological and environm ental value, contributing to aquifer recharge, water quality im provem e nt, flood m itigati on, and shoreline erosion protection. They

104also provide unique and extensive habitats for a wide variety of flora and fauna. Furtherm ore, wetlands play an im portant role in the gl obal cycles of nitrogen, sulfur, m e thane , and carbon dioxide (Mitsch and Gosselink, 1993).
The lower 48 states contained an estim a ted 100 m i llion acres (400,000 square kilom e ters) of wetlands in the m i d-1980s, an area about the size of California, am ong which freshwater wetlands are estim ated to m a ke up 80 to 95 pe rcent (Mitsch and Gosselink, 1993, Greene, 2000). Freshwater wetlands also cover extended areas of Alaska and Canada. Their proxim i ty to areas of hum an activity m a kes them susceptible to contam ination by petroleum hydrocarbons, via leakage, runoff or spill. Necessary m easures need to be taken to protect these ecosystem s since they are am ong the environm ents m o st sensit ive to oil and clean-up activities (Hayes et al ., 1995, 1997). Furtherm ore, there are reports that th e application of traditional oil spill cleanup techniques in wetland habitats caused m o re dam a ge than the oil itself (Baker et al. , 1993). Wh en looking for both inexpensive and environm entallyfriendly technologies for wetland cleanup and restoration, biorem ediation and phytorem ediati on have potential for being the m o st suitable options (Atlas, 1995; Cunningham et al, 1996).
The lim iting conditions for oil biodegradation in freshwater wetlands m ay be significantly different from m o st m arine environm ents. In te rm s of nutrient supply, freshwater wetlands can be divided into eutrophic wetlands (e.g., tidal m a rsh) and oligotrophic wetlands (e.g., cypress dom es (Mitsch and Gosselink, 1993). Many freshwater wetlands also exhibit a seasonal pattern of uptake and release of nutrients. During the growing season (i.e., late spring and summ e r), there is a high rate of nutrient uptake by vege tation from both the water and sedim e nts. And when higher plants die in the fall, a substantial portion of nutrients will be released to the water and sedim e nts. The am ount of inorganic nutrients or nutrients available for oil biodegradation also depends on m any other processes, such as nutrient m i neralization, denitrification, anaerobic release of phosphorus, and wetland hydrodynam i cs (Mitsch and Gosselink, 2000).
Wh en wetland soils are inundated with water, oxygen diffusion rates through the soil are drastically reduced. Available oxygen in the soil and in the interstitial water is quickly depleted through m e tabolism by aerobic organism s. Below a few centim e ters—and som e tim es only a few
m illim eters—of the soil surface, the environm ent becom e s anaerobic (Gam brell and Patrick, 1978). Wh en the m e tabolic dem and for oxygen by soil organism s exceeds that of supply, the redox potential in the soil drops and other ions (nitrate, m anganese, iron, sulfate , and carbon dioxide) are used as electron acceptors (Mitsch and Gosselink, 1993). Therefore, in freshwater wetland environm ents, petroleum biodegradation is likely to be lim ited by oxygen availability.
Because wetland sedim e nts are generally m o re fine-grained and, m o re im portantly, saturated with water, the extent of oil penetration is e xpected to be m u ch lower in freshwater wetlands than on a porous sandy m arine beach. In a m i croc osm study, Purandare (1999) investigated the penetration of weathered light Arabian crude oil in freshwater wetland sedim e nt under two dif f e rent water levels. For the case of low water level, where the sedim e nt was saturated but not covered with water, the oil was found to penetrate only about 2.5 cm in 16 weeks. For the case of high water level, where the water level was 10 cm above the sedim e nt surface, m o st of the oil was floating on the surface of the water and the penetration depth in the sedim e nt for som e settled and dissolved oil was also about 2.5cm duri ng this study. The depth of oil penetration in

105the case of the St. Lawrence River study was higher (about 9 cm ), due to the initial raking of the wetland sedim e nts af ter they were oiled. This depth is still m u ch lower than the depth of oil penetration that occurs in m arine sandy be aches (up to 30 cm , Gundlach, 1987) and cobble beaches (up to 1 m , W o lfe et al. , 1994).
Another im portant feature of wetlands is th at at least periodically, the land supports predom inantly hydrophytes, or plants “growing in wate r, soil, or on a substrate that is at least periodically deficient of oxygen as a result of excessive water content.” (Greene, 2000). These wetland plants m ay play im portant roles in oil biorem ediation and wetland restoration. On the one hand, they m ay be involved in degradati on, containm ent, and transfer of petroleum hydrocarbons from the soil to the atm o sphere (Frick, et al , 1999). On the other hand, these wetland plants m ay also com p ete with hydrocar bon-degrading m i croorganism s for nutrients.
5.5.2. Bioremediation strategies in freshw ater w e tlands Although the sam e decision-m a king and planning princi ples that were described earlier in this chapter for biorem ediation of m arine shorelin es should also apply to freshwater wetland environm ents, the feasible biorem ediation strategies are likely to be different due to the distinct characteristics of wetlands. The potential effectiven ess of different am endm ents is discussed as follows m a inly based on the findings of St . Lawrence River field study (Garcia-Blanco et al ., 2001b; Venosa et al. 2002; Lee et al ., 2001a).
Nutrient amendment Since nutrients could be lim ited in wetland sedi m e nts during the growing season in particular, addition of nutrients would seem to have som e pot ential for enhancing oil biodegradation in such an environm ent. However, the results from the St. Lawrence River field study (See Section 2.5) showed that no significant enhancem ent was obs erved in term s of the oil biodegradation following biostim ulation through addition of nutrien ts (either amm onium or nitrate). After 21 weeks, reduction of target parent and alkyl-subs tituted PAHs averaged 32% in all treatm e nts. Reduction of target alkanes was of sim ilar m a gn itude. The rem oval of PAHs in nutrient-am ended plots was only slightly better than natural attenua tion after 64 weeks of tr eatm ent. Oil analysis from the top 2 cm sedim e nt sam p les showed that the plots am ended with amm onium nitrate and with Scirpus pungens plants cut back dem onstrated a si gnificant enhancem ent in target hydrocarbon reduction over natural attenuation as well as all other treatm e nts. This suggests that biostim ulation m ay be effective only in the top layer of the soil, where aerobic conditions are greater, and when hydrocarbon-degrading m i croorga nism s do not have to com p ete for nutrients with the growing wetland plants. Another potential problem with resp ect to the use of biostim u lation in wetlands is that som e plant comm unities m ay be sensitive to nutrient additions. Repeated and excessive nutrient additions m ay alter the nature of the wetland eco system as indicated by the effects of chronic nutrient additions to the Everglades in Florida (Davis, 1994). Oxygen amendment Since oxygen has been found the m o st likely lim i ting factor in oil biodegradation in freshwater environm ents, oxygen am endm ent m ay be consider ed. However, an appropriate technology for

106increasing the oxygen concentration in such e nvironm ents, other than reliance on the wetland plants them selves to pum p oxygen down to the rhizosphere through the root system , has yet to be developed. Existing oxygen am endm ent technologies developed in terrestrial environm ents, such as tilling, f o rced aeration , and chem ical m e thods (See Section 4.4), are not likely to be costeffective for biorem ediation of freshwater we tlands since they often involve expensive and environm entally intrusive practices. During the St. Lawrence River field trial (Garcia-Blanco et al ., 2001b; Venosa et al. , 2002), after the first nutrient and oil applications, the top 1-2 cm surface soil in all plots was m anually raked using cast iron rakes. This was done to m i nim i ze lo ss of oil from the plots due to tidal action and to uniform ly incorporate the nutrients and the oil into the soil. However, the oil analysis results suggested that the tilling of surface soil m i ght have slowed the overall oil biodegradation rates by enhancing oil penetration deep into the anae robic sedim e nts. Based on these observations, surface tilling will not be an effective strate gy for increasing the oxygen concentration in freshwater wetlands (although this was not the in tent of the raking). The slightly better but statistically insignif i cant perf orm ance in both alkane and PAH degradation with addition of nitrate com p ared to amm onium after 64 weeks of treatm ent im plied that nitrate m ay have served as an alternative electron acceptor in enhanc ing oil biodegradation when oxygen was lim iting. However, the lim ited increase in biodegradation rate over natural attenuation would not warrant
a recomm e ndation to use nitrate as an oxygen am endm ent agent in such an environm ent. Phytoremediation Since plants cover wetlands at least periodicall y, the use of phytorem ediation becom e s a natural option for wetland cleanup and restoration. Phytor em ediation is em erging as a potentially viable technology for cleanup of soils contam inat ed with petroleum hydrocarbons (Frick et al. , 1999 See Section 4.3). However, this technique has not been used as a wetland oil spill counterm easure. Only lim ited studies have been carried out on the effectiveness of phytorem ediation in enhancing oil degradation in m arine shorelines and freshwater wetlands. Lin and Mendelssohn (1998) found in a greenhouse study that application of fertilizers in conjunction with the presence of transplants ( S. alterniflora and S. patens ) significantly enhanced oil degradation in a coastal wetland environm ent. In the case of freshwater wetlands, the St. Lawrence River study suggested that although appli cation of fertilizers in conjunction with the presence of a wetland plant (Scirpus pungens) m a y not significantly enhance oil degradation, it could enhance habitat recovery through the st im ulation of vigorous vegetative growth and reduction of sedim e nt toxicityand oil bioavailability (Lee et al ., 2001a). The effectiveness of oil phytorem ediation in freshwater wetland environm ents still requires further study.
Natural attenuation Natural attenuation has been defined as the reliance on natural processes to achieve site-specific
rem e dial objectives (USEPA, 1999b). This approa ch has been increasingly recognized as a possible viable option for oil spill cleanup with m o re understanding gained over the past decade about the advantages and disadvantages of activ e treatm e nt versus natural attenuation (Owens et al ., 1999). As indicated by Sell et al . (1995), the decision-m a ki ng should focus m o re on a preference for natural attenuation except when a large am ount of viscous oil is present, where natural rem oval will be slow, or when non-eco logical factors are of greater im portance.

107The St. Lawrence River Study dem onstrated that the availability of oxygen, not nutrients, is likely to be the lim iting factor for oil biodegr adation in freshwater wetlands. However, no feasible technique is currently available for increasing oxygen concentration under such an environm ent. As a result of this study, natural attenuation is recomm e nded as the m o st costeffective strategy for oil spill cleanup in freshwat er wetlands when the oil concentration is not high enough (e.g., less than 30 g/kg soil; Longpre et al., 1999) to destroy wetland vegetation.

108Chapter 6 GUIDELINES FOR ASSESSMENT OF FIELD RESULTS AND TERMINATION OF TREATMENT
After the treatm e nt is im plem ented according to the biorem ediation plan, the next or final step in an oil biorem ediation project is to assess the tr eatm e nt efficacy and term inate the biorem ediation action at appropriate treatm e nt endpoints. Key ques tions to be answered in this task include “what are the m easurem ents of oil biorem ediati on success?” and “how clean is clean?” or “when should biorem ediation efforts be term inated? ” . Actu ally, these issues should be dealt with during the biorem ediation planning stage, when proper treatm e nt objectives, strategies, and sam p ling protocols should be established. On the other hand, m o re definite answers to these questions can only be reached during the bior em ediation actions and based on the findings of com p rehensive
m onitoring program s. Cost-benefit analysis (e.g ., net environm ental benefit analysis, Baker, 1995&1999) and som e tim es political considerations should also be taken into account in this process, however, which are beyond the scope of this docum ent. From a technological point of view, the m easurem ents of biorem ediation su ccess and establishm ent of operational endpoints should be based on both the effi cacy of oil biodegradation and the evidence of ecotoxicity reduction and ecological recovery, each of which will be discussed in this chapter.
6.1 Assessment of Oil Biodegradation Efficacy
6.1.1 Verification of oil biodegradation
Evidence for the effectiveness of oil biorem ediation in term s of oil biodegradation should include: (1) faster disappearance of oil in treated areas than in untreated areas, and (2) a dem onstration that biodegradation was the m a in reason f o r the increased rate of oil disappearance. As described earlier in this docum ent, assessing the effectiveness of oil biorem ediation in oil spill sites is difficult due to the heterogeneous conditions of contam inated sites and lack of control over the oil distributi on. Nevertheless, the success of biorem ediation can be verified through well-designed m onitoring program s and proper data interpretation. Distinguishing biodegradation from abiotic loss Oil constituents can be lost from a shoreline by physical washout, dissolution, volatilization, and biodegradation. To dem onstrate the effectivene ss of a biorem ediation treatm e nt, biodegradation should be identified as the m a in m echanism for the increased rate of oil disappearance. As described in Chapter 3, to effectively distingui sh biodegradation from abiotic loss, specific oil
com ponents or analytes should be analyzed us ing GC/MS techniques and then these analytes should be norm a lized to a conserved biom arker, such as hopanes and chrysenes. This approach has been successfully used to distinguish betw een biodegradation and the physical or chem ical loss of oil in recent biorem ediation field studies (Bragg et al. , 1994; Venosa et al ., 1996; Lee et al ., 1997b). It should also be noted, however, thathopane norm a lization is m o st useful for reducing the variability associated with hete rogeneous oil distribution. The use of hopane
norm a lization to distinguish biodegradation from phys ical loss of oil is valid m o stly when losses due to dissolution and volatilization are negligible . Biodegradation can also be verif i ed as the
m a in rem oval m echanism by exam ining the relati onships between the degradation rates and the substrate structure such as the relative degr adation rates for hom ologous series of alkylated PAHs. These relationships, when used in conjuncti on with other oil analysis data, can be very

109useful in accurate assessm ent of the ex tent and progress of oil biodegradation (Venosa et al. , 1997a; W ang et al ., 1998; See section 3.3).
In addition to the dem onstration of oil bi odegradation based on chem ical analysis, the effectiveness of oil biorem ediation can also be verified by m onitoring the changes in growth and activity of oil degrading m i croorganism s. Grow th of hydrocarbon degrading bacteria can be determ ined by Most Probable Num b er (MPN) tec hniques, particularly the procedure proposed by Wr enn and Venosa (1996), which can separately enum erate aliphatic and arom atic hydrocarbondegrading bacteria and is sim p le enough fo r use in the field. Because m any viable
m i croorganism s are unculturable (Atlas & Bartha, 1987, Macnaughton et al ., 1999), the
em ergent culture-independent m o lecular technique s, such as the PLFA-DGGE techniques (See section 3.1.1.2), are becom i ng im portant tools to identify the diversity and com position of uncultivated m i crobial comm unities and to enum erate bacteria in m o re precise ways. Other usef ul tools in m onitoring biological activities include in-situ m e asurem ent of m i crobial CO 2 production by the use of respirom etric or radiorespirom e tric m e thods (Sugai et al ., 1997; Swannell et al ., 1997). Assessing treatment significance To show that a treatm e nt increases the rate of o il biodegradation, the concentrations of the target analytes (e.g., hopane norm a lized total resolvable alkanes or arom atics) should be significantly lower in treated than untreated areas within the tim e fram e of biorem ediation applications (e.g., several m onths to a year). Convincing dem onstrat ion of an increased rate of oil degradation requires taking a sufficient num ber of true replicate sam p les that are random ly interspersed throughout the sam p ling dom ain (see section 5.4). Th e field results should then be interpreted using proper statistical analysis (Venosa et al ., 1996), including analyzing field data using standard statistical m odels and analysis of variance (ANOVA) technique s, which are usually done at each sam p ling event. In addition to th e ANOVAs, the entire data set should be analyzed by non-linear regression analysis to estim ate the ra te of decline in biodegradable analytes for each treatm e nt over the entire course of the bior em ediation treatm e nt. As dem onstrated in the Delaware f i eld trial (Venosa et al ., 1996), the interpretation of treatm e nt effectiveness will be affected by reaction kinetics, background nutrient concentration, the variance of analytical results, the num ber of replicates, and statisti cal significance required to dem onstrate the differences between treated and control areas.
An exam ple of a likely treatm e nt effect under diffe rent kinetic conditions is shown in Figure 6.1. The perform ance of oil biodegradation expected on an oiled shoreline can be estim a ted based on the first-order kinetic m odels described in Secti on 2.5.3. In the Delaware field study, first-order biodegradation rate coefficients ranging between 0.026 and 0.056 day -1 f o r total resolvable alkanes and 0.021 to 0.031 day -1 for total resolvable arom atics were observed (Venosa et al ., 1996). These first-order biodegradation rate coe fficients are also a function of the nutrient concentration (See equation 2.4). Th e half-saturation concentration K nfor nitrate is approxim a tely 0.5 m g N/L (Boufadel et al. , unpublished). In this exam ple, the background nitrogen concentration is assum e d to be 0.1 m g N/L. The assum p tion of low background nutrient concentrations is reasonable since this is one of the prerequisites that biorem ediation actions should be selected. Two cases of oil biodegradation rates (i.e., 0.026 and 0.056 day -1 ) in the treated areas are exam ined, which covers both th e high (treated) and low (natural attenuation)

110ends of what was observed in the Delaware study (Venosa et al ., 1996). A theoretical control rate of 0.0093 day -1 was assum e d based on a background pore water nitrogen concentration of 0.1
m g /L and a Kn of 0.5 m g /L. This is the natural attenuation rate that would have occurred had the background nitrogen concentration been 0.1 m g /L rather than the actual 0.8 m g /L.

18 0
16 0
14 0
12 0
10 0

80
60
40
20

0 0 1 0 2 0 3 0 4 0 5 0 6 0 7 0 8 0 9 0 1 0 0 ktr e a t2 = 0

. 05 6 da y

-1 k contr o l = 0

. 00 93 day -1

k

tre a t1 =0

. 02 6 d a y

-1

Ti

m

e (d a y s ) Figure 6. 1 I n flu e nces o f biod egradation ra tes a on detectabi lity o f treatm ent effect

It can be seen from Figure 6.1 that differences between treated and control plots are prom inent very early in the project period when fi rst order decay coefficient was 0.056 day -1 . Di fferences were also evident when the decay rate was only 0.026 day -1 , but they were not as prom inent. Thus, had the background nitrogen concentration beencloser to 0.1 m g /L, the overall conclusion from the study would have been to recomm e nd bi orem ediation since a substantial enhancem ent would have been evident with nutrient addition. Th is would have been true even considering the
arom atic fraction, which had a lower decay rate but still higher than the theoretical control.
To determ ine the ef f ect of variance on the abi lity to detect dif f e rences between treatm ents, Figure 6.2 was developed. This figure was ba sed on the power analysis conducted by Venosa et al. (1996). Using 5 replicates and at a statisti cal power of 80% (i.e., the probability that significant differences between two or m o re treatm e nts are detected when they actually exist), the m i nim um detectible difference clearly increase s linearly with variance. Thus, if the variance doubles, the ability to detect a difference betweentreatm e nts lessens inversely. Figure 6.3 was developed to show how the m i nim um detectable diffe rence varies at a constant variance but at an increasing num ber of replicate plots. Obvious ly, the fewer the num ber of replicate plots established, the m o re dif f i cult it is to detect statistically signif i cant treatm e nt dif f e rences. These figures point out why it is so im portant to m i nim i ze variance by hopane norm a lization and increasing the num ber of replicate plots.TotalAlkanes(ng/nghopane)

111Figure 6.2 Minim um detectible treatm e nt difference as a function of variance.

150 140 130 120 110 100 90 80 70 60 50 40 30 20 10 0 0 2 4 6 8 1 0 1 2 number of replicatesminimum detectable difference, ng/ng hopane

Figure 6.3. Minim um detectible treatm e nt differen ce as a function of num ber of replicate plots.

1126.1.2 Assessment of physical loss
To better evaluate oil biorem ediation perform an ce, one should be able to distinguish between physical loss and biodegradative loss. This requi res a m a ss balance. However, little inform ation is available in this regard f o r m o st oil biorem ediation field studies and applications due to the lack of com p rehensive m onitoring program s, reliable m easurem ent tools, and proper data interpretation.
So f ar, the m o st com p lete m a ss balance of any m a jor oil spill was the Exxon Valdez incident. In the early 1990s, W o lfe et al. (1994) undertook a com p rehensive m onitoring program . They estim ated that about 3 years after the spill, a pproxim a tely 20% of the spilled oil had evaporated and undergone photolysis in the atm o sphere; appr oxim a tely 14% was recovered and disposed of; approxim a tely 2% rem a ined on intertidal shorelin es and 13% in subtidal zones. Approxim a tely 30% of the spilled oil was biodegraded in the water column , and nearly 20% was biodegraded on the shorelines. However, the paper did not provi de a m a ss balance for the 113 km of shorelines in Prince W illiam Sound where biorem ediation applications (nutrient additions) were perform e d.
A m e thodology was proposed by Venosa et al . (1996) to conduct a m a ss balance and to distinguish biodegradation from physical loss of oil in the Delaware field study. As m e ntioned earlier, spilled oils can be lost from a shorel ine by physical washout, dissolution, volatilization, and biodegradation. The m e thod assum e s that a nonbiodegradable com ponent of oil (nam ely, hopane) can be used to estim ate the first three lo ss rates, and that the actual biodegradation rate of an analyte can be estim a ted from the differen ce between its total loss rate and its physical loss rate. For this approach to be valid the physical wa shout rate of the oil m u st be dom inant, and the oil loss due to dissolution and volatilization m u st be negligible. Volatilization and dissolution
can cause som e preferential loss of oil com ponents particularly in the early stage of an oil spill (see Chapter 2). However, these factors are unlik ely to be im portant during the biorem ediation treatm e nt since biorem ediation is typically used as a polishing step after conventional
m echanical cleanup options have been applied and is often initiated weeks to m onths after an oil spill. W o lf e et al. (1994) reported that evaporation was no longer an im portant loss m echanism three m onths after the Exxon Valdez spill. Figure 6.4 shows the overall first order disappear ance of hopane and total extractable organic
m a tter (EOM) based on the results of the Delaware study. Over 90% of the spilled oil was
rem oved from the shoreline through physical wa shout based on the rate of hopane loss. The EOM first-order rate coefficient was higher th an the hopane disappearance rate. The difference in loss rates between hopane and EOM was attri buted to biodegradation because EOM includes both biodegradable and nonbiodegradable com ponents. However, because EOM and other total petroleum hydrocarbon (TPH) m easurem ents are not sensitive enough, no differences between biorem ediation treatm e nts and control could be dete rm ined using this approach in the study. This observation suggests that losses due to biorem edia tion m ay not be detectable using TPH analysis. This again dem onstrates that the success of biorem ediation should be judged by analyzing variables of biological significance, such as the reduction of concentrations of oil com ponents of ecological concern (e.g., PAHs), toxicity of th e oil, and ultim ately the ability to accelerate the recovery of the oil contam inated ecosystem

.

1131. 0

0. 8

0. 6

0. 4

0. 2

0. 0 0 1 0 2 0 3 0 4 0 5 0 6 0 7 0 8 0 9 0 1 0 0 H o pa ne : k = 0

. 02 5 EOM : k = 0. 03 4

P h

y s ic allo s s o f

o i l

T o tal o il l o s s B i od e g r a da t i o n

Ti m e , da y s

Fi gur e 6. 4 F a t eof oi ldi s a ppe a ra

nc e f ro m sh or e l i n eduri ng t h e D e l a wa r e s t ud y

Another im portant consideration in assessing physi cal loss of oil and biorem ediation success in a
m a rine shoreline is the effect of shoreline s ubstrate transport. Suidan and Wr enn (2001) found that substantial sand transport occurs over short tim e scales on m arine shorelines, particularly on high-energy beaches, although the net am ount of shor eline sedim e nt did not change significantly during the tracer study. Sedim e nt transport will affect oil transport since oiled sand will m ove with the bulk sand. As discussed in Section 3.2. 2, the physical redistribution of oil between the inside and outside of experim e nt al plots can affect interpretation of results from biorem ediation field studies that are conducted on shorelines contam inated by real oil spills. Studies of this type often involve treatm e nt of oiled sand in sm al l plots that are surrounded by large areas of untreated, contam inated beach. Because the treated sand will becom e m i xed with untreated sand from surrounding areas of the beach by wave-i nduced sedim e nt transport, the apparent effectiveness of the treatm e nt will be reduced . Hopane norm a lization cannot correct for this underestim ation of the treatm e nt effect since the hopane concentration inside the plots will not be af

f ected by sand transport. However, this ef f ect will be less im portant in cases of an actual biorem ediation application when the entire contam in ated shoreline is treated. The potential effect of

shoreline substrate transport on data inte rpretation in these cases, however, will be to overestim ate the treatm e nt effect due to the sand transport from large treated areas to sm all untreated control areas. Therefore, the contro l area set aside for assessing oil biorem ediation should be either large enough to reduce this effect or relatively isolated from the treated area toC/C0

114m i nim i ze the exchange of sand and nutrients betw een the two areas. The influence of shoreline substrate transport should always be taken into account in both the design of the sam p ling plan and the interpretation of data from the field, particularly for high-energy beaches.
6.1.3 Operational endpoints based on oil biodegradation
Biorem ediation endpoints for m any soil and groundw ater sites contam inated with petroleum products are often selected based on predeterm i ne d rem e diation target concentrations, such as 10-10,000 m g /kg TPH and 0.1-500 m g /kg BTEX, which are adopted by various regulatory agencies (King, et al ., 1998; Salanitro et al. , 1997). However, for reasons discussed previously in regard to the inadequacies of TPH analysis, su ch targets are not appropriate for protecting the environm ent. It is a good idea, however, for parties involved in a rem e diation project to have
som e m easurable endpoints for m anagem e nt and regulatory purposes. Based on existing experiences, the following criteria are suggested in determ ining biorem ediation endpoints with respect to oil biodegradation.

• Biorem ediation treatm e nt should be term inated wh en the extent of oil degradation tends to level off based on oil analysis results. Cost-ben efit analysis should be used in establishing target biorem ediation levels. It is unrealistic and uneconom ical to rem ove all traces of oil hydrocarbons using biorem ediation technologies.

• The concentrations of target oil analytes can al so be used as endpoint indicators, particularly when the treatm e nt is highly effective. Em pha sis should be given to those chem icals of environm ental concerns, such as PAHs. The target concentrations should be agreed upon in the treatm e nt plan and can be determ ined based on existing standards used for other environm ents (e.g., oil contam inated surface and subsurface soils, Bell et al. , 1994).

• The change in oil com position m ay also help to establish the biorem ediation endpoint. As oil becom e s m o re biodegraded, the fraction of le ss biodegradable com ponents (e.g., resins and asphaltenes) in the oil becom e enriched. Studies following the Exxon Valdez spill showed that oil biodegradation slowed substantially wh en the polar content of the North Slope crude oil reached 60-70% of the total m a ss (Bragg et al .,1994). Therefore, the polar fraction as a percent of the total oil m a ss rem a ining m ay potentially serve as an endpoint indicator although m o re research is still required to estab lish quantitative criteria. If rapid rem oval of the resin and asphaltene fractions of the oil is the desired endpoint, the only way to achieve this is by excavation and hauling to a contained or secure facility. That is because rem oval of these constituents in nature is known to occur only through dispersion and dilution. The rate of these processes can be very site specific and depend heavily on the type of substrate (i.e., cobble, sand, etc.) and wave energy.
6.2 Environmental Assessments
Biorem ediation is am ong the least intrusive of th e current operational physical and chem ical oil spill counterm easure options available. Ne vertheless, apprehension rem a ins about the environm ental im pact of biorem ediation agents released into the environm ent (Hoff, 1993; Holloway, 1991; Lee et al. , 1995a; Office of Technology Assessm ent, 1991). In addition to

115potential effects on wildlife and hum ans, there is concern that the by-products of enhanced oil biodegradation m ay be m o re toxic than th e parent com pounds. For general acceptance of biorem ediation as an oil spill counterm e asure, we m u st dem onstrate that it does not induce negative effects that suppress the rates of natural habitat recovery. Environm ental assessm ent should be an integral part of guidelines governi ng the application of biorem ediation treatm e nts to ensure protection of the environm ent. Both ecosystem structure and function m u st be considered in environm ental assessm ents. Ecosystem structure is studied by exam ining speci es abundance, biom ass, and diversity and other
com ponents at one point in tim e. Bioassessm ent fi eld surveys (Chapter 3.5.1) provide this basic inform ation on ecosystem structure. Ecosystem function describes the dynam i cs or changes in the system over tim e. Inform ation on ecosystem function is provided by the quantification of rates of biological processes like production, respiration, m i neralization, and nutrient regeneration. In addition, bioassays (Chapter 3.5.2) provide a m eans of quantifying the potential effects of toxicants on ecosystem function.
6.2.1 Operational guidance from environmental assessments for treatment application Controlled studies suggest that during rem e dial operations, optim al rates of degradation in sedim e nts can be achieved by sustaining elevated in terstitial nutrient levels that do not elicit an adverse effect (Lee et al ., 1993; Swannell et al. , 1996; Lee et al ., 1997b; Venosa et al ., 1996; Boufadel et al ., 1999b). Bioassessm ent field surveys can be used to guard against detrim ental effects such as the stim ulation of toxic al gal bloom s associated with eutrophication.
Environm ental assessm ents can also provide guidance for biorem ediation by pinpointing the optim al tim e for the onset of treatm e nt. It has becom e clear in num erous studies that treatm e nts have lim ited or no success as long as the residua l oil is retaining its m o st toxic com pounds, which are typically low m o lecular weight com pounds that are rem oved through natural weathering processes. In such a case, it is bett er to wait one to two w eeks before treatm e nt to allow f o r toxicity levels to decline, which can be delineated by tim e -series m onitoring of sedim e nt or water toxicity using bioassays. The results of toxicity tests have been used to explain the m ode of action and perform ance results of comm e rcial biorem ediation agents containing biostim ulation (nutrients) and bioaugm entation (bacterial inocula) properties. Lee et al. (1997b) observed that increases in m i crobial activity following treatm e nt are not necessarily correlated with toxicity reduction or habitat recovery. Natural attenuation was m o re effective because the biorem ediation agent inhibited the physical loss of residual oil from the sedim e nts. Standard bioassay test protocols are now being developed by regulatory agencies for toxicity evaluation of oil spill biorem ediation agents (Thom as et al. , 1995; Blenkinsopp et al. , 1995). Field application of biorem ediation products s hould be lim ited to those that have passed regulatory screening procedures for perform ance and toxicity. For a conservative approach, it is
recomm e nded that feasibility studies be conducted prior to full-scale operations. This can be
accom p lished by conducting contam inant biodegrad ability studies in the laboratory and
m onitoring the effectiveness of the proposed trea tm ent on several untreated but oiled shoreline

116segm ents. If the desired treatm ent endpoint (reduction of residual oil concentration or enhanced rate of habitat recovery) has been identif ied, ch em ical analysis and bioassays can be used to quantify the efficacy of biorem ediation treatm ents.
6.3 Case Study: Environmental Assessment of Bioremediation Treatments in a Tidal Freshw ater Marsh
Since environm ental assessm ent is a relatively ne w approach in evaluating the effectiveness of oil biorem ediation treatm e nts, a case study is pres ented in detail here to help spill responders better understand and conduct this type of assessm ent. To evaluate the efficacy of nutrient
am endm ent and phytorem ediation as biorem ediation strategies, a controlled oil spill field trial was recently conducted in a tidal freshwater m arsh on a shoreline of the St. Lawrence River, Canada to determ ine if nutrient enrichm e nt would enhance the rates of residual oil loss and habitat recovery (Blanco-Garcia et al. , 2001b; Venosa et al. , 2002; Lee et al. , 2001a). The experim e ntal design and biorem ediation perform an ce with respect to oil biodegradation has been reviewed in Chapter 2 (See Section 2.5.2). Envi ronm ental assessm ent of the extent of habitat recovery, which included a suite of bioassays f o r the identif i cation of possible detrim ental treatm e nt effects (e.g. toxicity of the biorem ed iation agent or oil degradation by-products), is described as follows.
6.3.1 Alterations in ecosystem structure Vegetative recovery of the predom inant plant species ( Scirpus pungens ) was m o nitored by determ ining changes in species com p osition (predom inant S. pungens and secondary Eleocharis palustris ) , t o t a lb i o m a s s , h e i ght , a n d p e r c e n t c o v e r
. During the first growing season S. pungens tolerated the experim e ntal oil concentrations, but suffered oil-induced growth inhibition (Figure 6.5), which is consistent with the re sults of recent greenhouse studies (Longpré et al. , 1999). The ef f ects of growth inhibition by oil were still evident within natural attenuation plots at W eek 65. The enhanced recovery of the vegetation in nutrient am ended plots during the second field season could not be attributed to the concentrat ion of nutrients in the porewater, as they had dim i nished to background levels before the previous winter. Analysis of nitrogen content within the roots of plants suggested that the abundant growth was attributable to the recycling of organic nitrogen stored by these plants over the first season.
W ith the reduction of residual oil concentrations (prim arily by physical rem oval) and regrowth of the predom inant vegetative species, extens ive recolonization by indigenous invertebrate species such as the my stery snail ( V. georgianus ) was also observed during the second field season in all experim e ntal plots (W eek 45).

117Na t

. Att n . NH 4 + NO 3 - NH 4 + Unoi led 0 20 40 60 80 100 120 140

Figure 6.5 Average height of th e predom inant plant species ( Scirpus pungens ) at W eek 15 in three oiled treatm e nts (Nat. Attn., NH 4 + , NO 3 - ) and the unoiled but fertilized (NH 4 + ) treatm e nt.
6.3.2 Alterations in ecosystem function
The environm ental im pact of contam inants in aquatic ecosystem s cannot be assessed accurately with a single species bioassay, which cannot repres ent the range of sensitivity of all biota. address this issue, a test battery with orga nism s from several trophic levels was used: (1)
m i crobial response; (2) Microtox solid phase; (3) algal solid phase; (4) cladoceran survival; (5)
am phipod survival; (6) gastropod survival/histopathology; and (7) acute and chronic effects in fi sh. 6.3.2.1 Microbial response Concentrations of bacteria are up to seven orde rs of m a gnitude higher in the surface sedim e nts than in the water column
. m i crobial activity m a kes sedim e nts the m o st active site for transform a tions of organic car bon, nitrogen, phosphorus, m a gnesium , and sulfur. If the processes of decom position, m i neraliza tion, and nutrient regeneration are disturbed by contam inants, the nature of the ecosystem will be changed. portance of these and other processes concentrated in the sedim e nts, relatively few assessm ents have focused on functional changes attributable to sedim e nt-associated contam inants.
Detailed studies on the m i crobial response to bi orem ediation treatm ents within experim e ntal enclosures were conducted by Greer et al. (2000). showed a slight increase during the f i rst 4 weeks f o llowing oil and f e rtilizer application. increase was clearly due to the fertilizer, as ev ident from the contrasting population densities in untreated areas which rem a ined relatively unc hanged throughout the m onitoring period. Patterns of i neralization nd
m i croorganism s can be used as an indication of in-situ biodegradation of petroleum (Braddock et al. , 1996). i crobial populations dem onstrat ed a rapid and sustained increase in naphthalene m i neralization activity in the plots that were both oiled and f e rtilized. Activity was
som e what lower in unfertilized/oiled and ferti lized-only plots. i neralization To

This concentration of

Despite the im

The viable bacterial population density The

m hydrocarbon a activity hydrocarbon-degrading of distribution
The m
Hexadecane m

118activity increased in response to fertilizer a pplication, especially amm onium nitrate, in
com p arison to sodium nitrate: activity in the unfertilized/oiled plots and unoiled reference control areas rem a ined relatively low. Laborator y assays to m onitor various pathways in the nitrogen cycle (nitrif ication, denitrif ication, ni trogen fixation) have been developed (Pritchard and Bourquin, 1985) and can be used to assess ch anges in function due to contam inant stress. Field and laboratory evaluation of nitrogen m e tabolism indicated significant denitrification activity in sedim e nts f o llowing f e rtilizer applicati on, which was not adversel y affected by oiling. In contrast to the results of chem ical anal ysis that showed no treatm ent effect on oil biodegradation rates, the results dem onstrated thatthe application of f

e rtilizers stim ulated the activities of indigenous hydrocarbon degrading and de nitrifying bacteria, and the presence of oil did not have a detrim ental effect on these activities. 6.3.2.2 Microtox solid phase test In the M i c r o t o x ®S o l i d P h a s e T e s t (Microbics Corporation 1995), t h e b a c t e r i u m , Vibrio fisheri , was exposed to oiled sedim e nt. A s i gn i f i c a n

t d e c r e a s e i n b i o l u m i n e s c e n c e r e l a t i v e t o w a t e r - o n l y controls was indicative of sedim e nt toxicity. Toxic ity levels were calculated as the concentration of
sam p le that would result in a 50% reduction in lum i nescence (‘effective concentration,’ EC 50 ). To account for interferences from differences in sam p legrain size distribution, turbidity, and to a lesser extent, color of the sam p le dilutions, sam p le test results were com p ared with results f r om unoiled sedim e nts from the im m e diate study area.
O il toxicity was evident on com p arison of oile d with unoiled plots (Figure 6.6). Wh ile the f e rtilized plots showed a trend towards a reduction in toxicity, there was no such evidence in the natural attenuation plots over 65 weeks. Major tr eatm ent effects were not observed in the first field season. However, in the second field season th e relative toxicity levels in nutrient am ended plots were sim ilar to values of the unoiled controls. 6.3.2.3 Algal solid phase assay An algal solid-phase assay (ASPA) was used to a ssess the toxicity of sedim e nts recovered from the experim e ntal test plots (Blaise and Ménard, 1998). The endpoint for this assay is based on the concentration of sedim e nt causing 50% inhibition (IC 50 ) of esterase enzym e activity in Selenastrum capricornutum due to toxicants. This test is an excellent biom arker for environm ental assessm ents, as esterases are a key group of ubiquitous enzym e s found in both plants and anim als (Dorsey et al. 1989 ; Gala and Giesy 1990). Results of the algal solid-phase assay (Figure 6. 7) showed that there was no toxicity in the unoiled reference plots throughout the course of the experim e nt (IC 50 > 7%), and that nutrient additions to unoiled sedim e nt appear to cause no detrim ental effects. For the first 21 weeks, elevated levels of toxicity were observed in all the oiled plots. Significant reduction in sedim e nt toxicity was not observed until the second field se ason. At this point, a m a rked decrease in toxicity was observed for the oiled plots am ende d with nutrients (either amm onium nitrate or sodium nitrate). Nutrient am ended plots were deem ed non toxic by W eek 65. Wh ile nutrient treatm e nts were term inated at the end of the firs t field season, treatm e nts resulted in significant positive effects that were not observed until the second field season. This lag suggests that toxicity reduction m ay be correlated with enhanced vegetative growth associated with nutrient enrichm e nt.

119

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. NH 4 + NH 4 + NO 3 U n o ile d A

ttn

. Cut N H 4 + 

## T r eat m ent s Figure 6.6. Tim e -series changes in sedim e nt toxicity as quantified by the Microtox  Solid Phase Test. Error bars = 1 standard deviation

8EC50(ppt)01246812162149650124681216214965012468121621496501246812162149650124681216214965

7

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### not t o xic

0 1 6 12 2 1 4 9 6 5 0 1 6 1 2 21 49 65 0 1 6 1 2 21 49 65 0 1 6 1 2 2 1 4 9 6 5 0 1 6 1 2 2 1 4 9 65

- N a t. A t tn. NH 4 + C u t P l a n ts NH 4 + NO 3 Uno i led NH 4 + 

### T r ea t m en t s Figure 6.7. Algal solid phase assay (ASPA) as IC 50 for sedim e nt sam p les collected at W eeks 0, 1, 6, 12, 21, 49 and 65. Bars represent m eans ± standard deviation.

1206.3.2.4 Cladoceran survival test The cladoceran Daphnia magna (‘ water flea’) was used to assess the toxicity of elutriates from sedim e nt sam p les (Environm ent Canada, 1990). Treatm ent of unoiled plots with amm onium nitrate and super triple phosphate appeared to cause negligible im pacts on the survival of D. magna (Figure 6.8a). In contrast, imm e diate, butlim ited, toxic responses were observed in all oiled plots. Amm onium nitrate and triple super phosphate additions appeared to reduce residual oil toxicity to background levels by W e ek 2. Am endm ents with sodium nitrate appeared less ef

f ective. Since the sensitivity of this assay appeared to be lim ited — the ef f ect of residual oil was deem ed negligible by W eek 6 — this assay was excluded from the second field season. 6.3.2.5 Amphipod survival test The Am phipod Test m easured the effects of se dim e nt sam p les on survival of s e d i m e n t

- d w e lling Hyalella azteca neonates, 2-9 days old (Environm ent Canada, 1997). Both the m ean percent survival and the m ean weight of anim als in eachtreatm ent were com p ared with m ean percent survival and m ean weight of am phipods in refe rence control sedim e nts to determ ine if the treatm e nts caused a significant decrease in organism survival or growth. Hyalella azteca m o rtality (Figure 6.8b) was a m o re sensitive bioassay endpoint (i.e. higher response) than Daphnia m o rtality (Figure 6.8a). This m ay be due to species differences and the fact that the Am phipod Survival Test is a direct -contact sedim e nt test. Oil derived toxicants within the sedim e nt m ay not have been effectiv ely transferred to the elutriate used in the cladoceran test. Furtherm ore, in a laboratory test, the uptake of sedim e nt-associated anthracene by the freshwater am phipod Hyalella azteca was reported to occur at a rate m u ch higher than predicted. It was concluded that selective feeding of H. azteca on sm aller particles results in a diet of fines containing the highe st organic m a tter concentration and, hence, contam inant levels (Landrum and Scavis, 1983). H. azteca m o rtality was consistently lower in the unoiled nutrient
am ended plot. High m o rtality was observed during th e f i rst week in all oiled plots. The presence or absence of plants appeared to have no significant effect. At the latter part of the first field season there appeared to be a pronounced increase in sedim e nt toxicity in the oiled plots am ended with prilled amm onium nitrate. The sensitivity of H. azteca to amm onia (LC 50 = 14.9 m g -N/L) was verified in th e laboratory. Many sam p les, notably W eeks 12 and 21, had overlying water amm onia leve ls exceeding the established toxicity lim its, with the highest values corresponding to the oile d plots am ended with amm onium nitrate. The fertilized unoiled site had lower elevated amm onia levels. This and a possible synergistic effect between oil and elevated NH 4 + concentrations m ay explain th e observation of little or no toxic response observed in the fertilized unoiled plots (Figure 6.8b). In response to a reduction in residual oil concentrations, percentage m o rtality was near background levels by the second field
season. The results of this bioassay suggested that the experim e ntal treatm ents offered little advantage on an operational scale.

12150

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### b)

0126 1 2 2 1 4 9 6 5 0 126 1 2 2 1 4 9 6 5 01 26 1 2 2 1 4 9 6 5 01 2 6 1 2 2 1 4 9 6 5 012 6 1 2 2 1 4 9 6 5

- Nat

. NH 4 + NH 4 + NO 3 Unoiled At t n . Cu t N H 4 + T r eatment s Figure 6.8 Tim e -series changes in sedim e nt toxic ity as quantified by: a) Cladoceran Survival Test; b) Am phipod Survival Test. Error bars = 1 standard deviation 6.3.2.6 Gastropod survival/histopathology Although m any organism s have been used as sentinels or biom onitors of environm ental contam inants (LeBlanc and Bain, 1997), there is s till a need to identify and exploit alternative species that are sensitive and am enable to ecot oxicological testing. Molluscs are abundant and widely distributed and their use as in situ biom onitors has been on the rise (Lagadic and Caquet, 1998). In this study (Lee et al. , 2001b), gastropod survival and histopathology assays were conducted with the my stery snail, Viviparus georgianus . They were specifically selected for use
as an in situ biom onitor as they feed on sedim e nt de tritus, algae and decaying organic m a tter within the wetland. Snails (n = 50/treatm e nt/sam pling tim e) were caged within 20 × 20 × 22 cm open m e sh polypropylene baskets m oored to the sedi m e nt surface of experim e ntal plots, and in designated ‘untreated control’ areas within the vici nity of the plots. Cages were recovered after 30, 60 and 90 days of exposure to evaluate eff ects on survival at the end of the second field%Mortality%Mortality 

### a)

122season (W eek 65). Healthy snails were also exposed for a 30 day period to test sedim e nts recovered from the plots, under laboratory conditions, to determ ine survival rates. Significant growth (p < 0.001) of the snails was observed during the study, but no significant differences (p > 0.05) were observed in tissue weight, shell size or shell thickness between the experim e ntal treatm e nts. Under adverse conditions V. georgianus is known to retract into its shell and show no m o tility for prolonged periods, so anim al vitality was assessed by the presence of an operculum . No surviving snails were f ound in any of the continuous exposure cages after three m onths (Figure 6.9). To factor out the in fluence of stress from the caging of anim als for extended periods, data at each sam p ling event were norm a lized to unoiled control plots adjacent to the test blocks, which received no oil or nutrient am endm ents (4%, 47% and 100% m o rtality at 30, 60 and 90 days, respectively). The only oi led treatm e nt that gave a higher percentage survival after 30 days exposure was the amm onium nitrate am ended plots with intact plants. The other treatm e nts were m o re toxic than the na tural attenuation plots. Elevated nutrient concentrations for 60 days did not reduce th e toxic effects of the oiled sedim e nts to V. georgianus. Indeed, exposure to the treatm e nt probably exacerbated stress.


### 100

75
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0 NH 4 NH 4 + NO 3 - Unoiled, NH 4 + Unoiled Nat u r a l Att enu a t io n

+ , C u t Pl an t s



### 30 60 E x posu re T i m e ( d a y s) 90
Figure 6.9 Percent survival of the my stery snail, Viviparus georgianus , following exposure to test treatm e nts for 30, 60 and 90 days. Resu lts norm a lized to survival in unoiled sedim e nt without nutrient am endm ent.

Toxicity of nitrogenous com pounds at high levels has been noted with various organism s. Am ong invertebrates, amm onia toxicity was a ssessed with a freshwater snail (Hickey and Vickers 1994). Acute values were derived for the snail to be 0.15 g/m amm onia com p ared to the EPA value of 0.52 for salm onids, and they found that the snail was m o re sensitive than the
norm a lly accepted sensitive species such as m ayfli es and stoneflies. After 30 days of continuous exposure, the highest m o rtality was reported in the fertilized plots with plants cut back. Direct observations suggested that these anim als suffe red from harsh conditions in the absence of natural ground cover. All treatm e nts containing nutrients with or without oil caused higher
m o rtalities than no oil or oil alone. Laboratory exposures using sedim e nt recovered from the experim e ntal plots at W eek 65 (which had less oil, and of lower bioavailability) showed that the application of treatm e nts in the first season ha d insignificant long-term effects (Figure 6.10).

123Based on the results of the am phipod and snail bioassays, biorem ediation using nutrients in situ will need to take the above into consideration.

100

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. NH 4 + NH 4 + NO 3 - U noil e d At tn. Cut NH 4 + Figure 6.10 Percent survival of the my stery snail, Viviparus georgianus under laboratory conditions following 30 days of exposure to sedi m e nts sam p led at W eek 65. Error bars = 1 standard deviation

Wh en organism s are exposed to xenobiotics, cel lular changes have been observed to occur. Evaluation of histological change s can provide im portant inform ation as to the stress of the organism and m ode of action of pollutants. Specim ens of live snails were preserved for histological analysis. All tissues showed degenerative changes over tim e in snails exposed to fertilizers. Pathological changes were m o st obvious in epithelial cells lining such tissues as gills, intestine and digestive gland, but som e degenerative changes were also observed in other organs such as gonads. The m o st dram atic changes were observed in snails from plots treated with sodium nitrate (Figure 6.11). Effects of treatm e nt on reproductive success of V. georgianus were noted. Gravid fem a les contained on average 10 young (range 3 to 15) visible with the naked eye. At 1-week post oil exposure, gravid fem a les s howed som e degenerating em bryos in all but the control treatm e nts. This was also true for som e of the 1-m onth post exposure gravid fem a les, but
num ber of gravid fem a les with m acroscopic em bryos decreased. 6.3.2.7 Acute and chronic effects on fish Experience with the Exxon Valdez dem onstrated that oil deposition in shoals where pink salm on ( Oncorhynchus gorbuscha ) and Pacific herring ( Clupea pallasi ) spawn, caused blue sac disease (BSD) of newly hatched larvae (Hose et al ., 1996; Marty et al. , 1997). BSD is characterized by yolk sac and pericardial edem a, hem o rrhaging, deform ities, and induction of m i xed function oxygenase enzym e s (Marty et al. , 1997; Billiard et al. , 1999; Fragoso et al. , 1998). As a consequence, there is considerable interest in the application of biorem ediation strategies that accelerate the rem oval of oil from intertidal beaches, which provide the nursery habitat for m a ny fish species. The success of these technologies should be judged not sim p ly on how quickly oil disappears, but on a dem onstrated reduction in ri sk, i.e. on how quickly toxicity (hazard) and exposure to oil are reduced.% Survival

124

### a

b Figure 6.11. Representative sections through th e intestine of Viviparus showing norm a l villi seen in control snails (a) and degenerated villus with hem o cytic inf iltration seen in snails f r om the sodium nitrate treatm e nt after 1 m onth exposure (b). Bar = 75 μm

.
Establishing the exposure of fish to oil is difficult

. Wh ile chem ical m easures of oil in sedim e nts, water and tissues are routine, there is no guarant ee that fish accum u late oil or its com ponents equally or in proportion to environm ental concentrations. Further, m any of the com ponents of oil (e.g. alkanes, polynuclear arom atic hydrocar bons (PAH)), are m e tabolized by fish, so that chem ical analyses of fish tissues m ay not represent the true dose or dose rate. Assessing hazard is equally difficult. Oil is a com p lex m i xture, and m any of its com ponents have dif f e rent m odes of toxicity. At acutely leth al concentrations, m o rtality is rapid and is likely due to narcosis caused by m onoarom a tics (benzenes, toluenes, xylenes, etc.) and alkanes. At lower concentrations (Billiard et al. , 1999), chronic toxicity (BSD) m a y be linked to the concentrations of alkyl PAH such as 7- isopropyl-1-m ethylphenanthrene (retene, a C-4 phenanthrene). This m e chanism m ay involve m e tabolism of alkyl PAH to m o re toxic form s by CYP1A enzym e s. Toxicity and rate of excreti on of phenanthrene and retene can be m odulated by inhibiting or inducing CYP1A activity (Hawkins et al. , 2000). Delayed responses, such as the long-term onset of cancer, m ay also follow brief exposures to pro-carcinogens such as benzo(a)pyrene (BaP). In this case, the m echanism involves oxygenation of BaP by CYP1A enzym e s to carcinogenic diols and epoxides, whic h cause genotoxicity by form ing adducts with bases of DNA (Varanasi et al. , 1989). In this case study, bioassays of CYP1A induction in juvenile rainbow trout ( Oncorhynchus mykiss) , BSD in em bryo-larval stages (ELS) of trout, and reproductive developm ental studies

125with Japanese m e daka ( Oryzias latipes ) were used to evaluate the presence, bioavailability, and toxicity of PAH (and parent m i xture of o il) in contam inated beach sedim e nts (Hodson et al. , 2001). Changes in bioavailability and toxicity with tim e were m onitored by analysis of tim eseries sam p les. The key to sedim e nt assessm ent is bioavailability. Although sedim e nts m i ght contain relatively high concentrations of toxic com pounds, this c ondition does not necessarily lead to adverse effects on organism s living in the sedim e nts (Payne et al. , 1988). The only m eans of m easuring bioavailability is by m easuring or determ ining biological response. Such testing has often involved m easures of bioaccum u lation (the ability of an organism to accum u late contam inants in tissues. However, because bioaccum u lation is a phenom enon, not an effect (and can be relatively expensive due to costly chem ical anal yses), em phasis has shifted towards indicative endpoints that are based on sedim e nt toxicity tests, which are effects-based and relatively inexpensive. Bioavailability was assessed by the extent of CY P1A induction in fingerling trout after a 4 d exposure to sedim e nts. S-9 fr actions were prepared from liver hom ogenates and activity of ethoxyresorufin-o-deethylase (EROD, CYP1A enzy m e ) was m e asured by a kinetic m i croplate fluorescence m e thod following the protocols outlined by Hodson et al. (1996). Each bioassay included negative controls (water only), unoiled sedim e nt controls, and positive water controls (fish exposed to ß-naphthoflavone, a m odel inducer). Results indicated that one day after oiling, EROD activity of fish exposed to oiled sedim e nts was on average 25-fold higher than that in the control sedim e nts (range = 13-35-fold). Over a 17 m onth period, th ere was about an 80% decline in induction potency by oiled sedim e nts sam p led from the beach. The decline in EROD activity paralleled declines in total hydrocarbons , total PAH, and total alkyl-PAH, m e asured by GC/MS with correlation coefficients betweenlog of EROD activity and log of hydrocarbon concentrations of 0.96 or higher (Hodson et al. , 2001). Blue sac disease of trout was assessed by exposi ng 50 eyed-eggs (about 15 d post fertilization) to sedim e nts until they had hatched, resorbed the m a jority of their yolk sac, and begun to swim up (Zam bon et al. , 2000). After 32 d of exposure, larvae were rem oved and the % survival and prevalence of the sym p tom s of BSD (edem a , hem o rrhaging, def o rm ities) were m e asured (Guiney et al. , 1997). Results showed that trout em bryos and larvae exposed to oiled sedim e nts from Ste. Croix exhibited a low prevalence of sym p tom s of BSD, although rates were higher than observed for fish exposed to the referenc e sedim e nts. All oil-exposed fish exam ined histologically dem onstrated intense staining for CYP1A protein, indicating a significant exposure to CYP1A-inducing com pounds such as PAH. These prelim inary results suggested that sym p tom s of BSD were m o re frequent in fi sh exposed to oiled than to un-oiled sedim e nts, indicating a risk to early life stages of species that spawn on tidal beaches. Medaka larvae exposed to oiled sedim e nts for 90 days (m aturity) experienced a higher m o rtality rate (42.5 ± 4.2%) than those exposed to un-oiled sedim e nts (9.1 ± 7.4). The growth of surviving
m e daka in the oiled sedim e nt treatm e nts was im paired (Figure 6.12), as shown by their sm aller size com p ared to controls (20.6 ± 1.3 vs 16.4 ± 3.3 mm ) . In addition, nearly all m e daka from the oiled sedim e nt treatm e nts had deform ed or m i ssing fins. Prelim inary histological exam ination of

126fixed m e daka indicated a degenerative liver lesion, spongiosis hepatis , in >95% of the fish from the oiled-sedim e nts, but not in controls. There was also a low incidence of m a le m e daka (<10%) with intersex gonads (testis-ova) in fish exposed to oiled sedim e nts.

24

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12 U n - o ile d O i l e d S e d i m e n t W a t e r s e d i m e n t ( E ) ( P o o l e d A- D) ±S t d

. D e v .

±S t d

. E r r . Me a n a

a

b

Figure 6.12. Effect of exposure to Ste. Croix oiled sedim e nt (Oct. 1999) on growth of m e daka. Treatm e nts sharing the sam e letters are not statistically different.

The laboratory fish bioassay represents a ‘worst -case’ scenario as the fish could not avoid exposure. Furtherm ore, m i xing of the sedim e nts in the test cham bers m i ght have disrupted the top-m o st sedim e nt layer (that m i ght have been depleted of oil in the field by weathering) and exposed m o re oily layers of sedim e nt. The ratio of water to sedim e nt is also fixed, which is very different from the situation in well-flushed tidal beaches. Finally, the test organism s were not beach spawning species of fish such as sm elt, capelin, and herring.
Sem i -perm eable m e m b rane devices (SPMDs) were used to assess in-situ changes in the bioavailability and toxicity of residual oil com ponents following treatm e nt. These passive in-situ sam p lers, developed by Huckins et al. (1990), contain purified triolein, a substance that constitutes a m a jor f r action of the neutral lipid of f i sh. Wh en imm e rsed in water, SPMDs absorb non-ionic, organic chem icals having a log K ow >1, a size <1 Å, a m o lecular weight of about 600 or less, and possibly neutral organo-m e tal com p lexe s. These characteristics correspond to that of known m i xed-function oxygenase (MFO) i nducing com pounds, including PAHs. The operational advantage of SPMDs is that they can be deployed within the environm ent to provide an integrated sam p le over tim e. This charact eristic m a kes the assay highly advantageous for field use. In term s of ecological relevance, th e diffusion of dissolved neutral organic chem icals into the triolein through the pores of the polye thylene m e m b rane within SPMDs sim u lates the diffusion of com pounds across a live fish gill m e m b rane. The lipid can be analyzed by traditional chem ical techniques to provide a list of chem icals absorbed, their concentrations inTotalLength(mm)

127the SPMDs, and, by back calculation, their concentr ations in water or sedim e nt. In this study SPMDs were used as concentrating device s for biological testing with the Microtox  Assay. SPMD units (12.5 cm ) enclosed in protective cases were deployed at the water sedim e nt interface at predeterm i ned intervals and recovere d for analysis after one week of exposure. SPMD concentrated sam p les were diluted with organic solvents for toxicological analysis by Microtox. The results showed that by 4 weeks, th e toxicity of sedim e nts within the oiled plots had declined to the levels of the unoiled but f e rtilized control as the result of a reduction in bioavailability (Johnson et al. , 2000).
6.4 Ecotoxicological Tests for Risk Assessment The application of in-situ biorem ediation operations is not expected to generate a large volum e of waste m a terials like ex-situ operations do. If the program is effective, any residual hydrocarbons will elicit little or no biological e ffect due to physical, chem ical, and biological processes that reduce their bioavailability. As i llustrated in the case study, sedim e nt quality can be assessed by a num ber of m e thods that tend to fit into one of five categories: sedim e nt chem istry, sedim e nt toxicity, comm unity structure, tissue chem istry, and pathology. Ideally, all five com ponents would be utilized to asse ss sedim e nt quality. However, in reality, environm ental m anagers are faced with lim itati ons in both resources and tim e
. They m u st optim ize use of their resources by selecting the inf o rm ation that will have the greatest utility. Thus, m onitoring program s should be focused on the m easurem ent of variables that allow quantification of treatm e nt success against a pre- defined endpoint. Project coordinators m u st strive to strike a balance between level of effort and type or quality of inform ation needed to
m a ke effective decisions.
Wh ile detrim ental effects have not been linked to the application of biorem ediation strategies based on nutrient enrichm e nt in actual spill response operations (Mearns et al. , 1997; Prince, 1993), the results of recent field trials clearly dem onstrate that the possibility exists. In the case study presented (Section 6.3), im proper applicati on of biorem ediation agents (the addition of excess f e rtilizer) was detrim ental to the environm en t. For exam ple, synergistic ef f ects between
amm onia and the test oil were observed in the Am phipod Survival Test. However, the nutrient additions had no effect on the unicellular algal species, Selenastrum capricornutum , and actually enhanced the growth and productivity of the dom inant plant species within the oiled plots relative to those that received no treatm e nt. Rapid recovery of vegetation is critical within an im pacted wetland for erosion control. Depe nding on the desired endpoint of the rem e dial operation, a balance m u st be m a de am ong th e m any positive (e.g. enhanced recovery of vegetation) and negative effects of treatm e nt (e .g. changes in productivity, species com position, and diversity am ong the rem a ining wetland plants).
W ith refinem e nt, bioassays can be used in oil spill operations to provide real-tim e guidance to the treatm e nt operations (e.g. determ ining the optim al nutrient concentration that does not elicit a detrim ental effect), to verify the success of count erm easures and to quantify the extent of habitat recovery. The dem onstration of species dependentresponses in previous investigations and the current case study suggest that future envir onm ental risk assessm ents be based on a m u ltispecies, m u lti-trophic level test battery approach. The results of the ecotoxicological tests should

128be used to build an ecological risk assessm ent — an estim a te of the probability of harm to the aquatic environm ent derived from the synthesis of results of separate exposure and effects
com ponents in a scientific m anner (Gentile et al. , 1989). It is often stated that the objective of oil spill counterm easures are to return an im pacted site to its imm e diate pre-spill condition. This is an unrealistic goal as the environm ent is a dynam i c rather than a static system

. On an operational scale, our goal should be to return the structure and function of an ecosystem to within the lim its of pre-defined, acceptable criteria.
6.5 Ecotoxicological Tests to Identify Operational Endpoints The effectiveness of oil spill counterm easures ultim ately m u st be judged by their ability to reduce injury to aquatic life. The worse case scenario would be the use of a response m e thod that is not effective in reducing exposure and incr eases injury to aquatic life. In biorem ediation operations the application of ecotoxicological m onitoring protocols m a y be used to verify the efficacy for toxicity reduction over that of no treatm ent.
Biorem ediation treatm e nts should be term inated wh en it is deem ed that: (1) treatm ents offer no operational advantage over natural recovery, (2) the contam inant concentrations and toxicity values are reduced to acceptable levels, or (3) de trim ental effects from the treatm e nt strategy are identified. Cost-benefit analysis should be consid ered in the decision of the acceptable level. It is futile to expect biorem ediation techniques to rem ove all traces of residual hydrocarbons. In term s of ecological relevance, declaration of habi tat recovery can be m a de when toxicity lim its are within regulatory guidelines and there is evid ence f o r the return of the original comm unity structure (Lee et al. , 1995b; Mearns et al. , 1995).

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